Denitrification As an Adaptive Trait in Soil and Groundwater Bacteria

Home , Aerobic denitrification, Denitrification

Denitrification as an adaptive trait in soil and groundwater bacteria

Christer Bergwall

NO;I I

Department of Ecology Chemical Ecology & Ecotoxicology Lund University, Sweden Lund 1997 Denitrification as an adaptive trait in soil and groundwater bacteria

Christer Bergwall

.' t , '. B.Sc., W1

Ph.D. Thesis Lund 1997

I. A doctoral thesis at a university in Sweden is produced either as a monograph or as a collection of papers. In the latter case, the introductory part constitutes the formal thesis, which summarises the accompanying papers. These have either already been published or are manuscripts at various stages (inpress, submitted or in manuscript).

01997 Christer Bergwall ISBN 91-7105-0884 SE-LUNBDS/NBKE-97/1012+94pp DISCLAIMER

Portions of this document may be illegible electronic image products. Images are produced from the best available original document.

I, Oganiration Documentname LuNDuNIVERsrn DOCTORAL DJSSERTATION ~~~ Department of Ecology Date of September 30,1997 Chemical Ecology & Ecotoxicology S-223 62 Lund, Sweden CODEN: SE-LUNBDS/NBKE-97/1012+94pp Authol(r) spolmrillg oganiration Christer Bergwall

riandsnbtitle Denitrification as an adaptive trait in soil and groundwater bacteria

Abstract The focus of this thesis is on selection and adaptation processes in bacteria with emphasis on denitrifying bacteria in groundwater. Other nitrogen transformation processes such as dissimilatory nitrate reduction to ammonium (nitrate ammonification) and nitrification of forest soil bacteria are briefly discussed. Microcosms with sterile sediment and groundwater were inoculated with single denitrifying strains isolated from three groundwateraquifers, two which are agricultural aquifers (in siru NOi-N was 24.1 and 35.2 mg r') and the third which is a pristine lake water infiltration aquifer (in situ NO;-N was 6.3 mg F'). The average denitrification activity for strains from the nitrate contaminated sites were twice as high as the activity of the strains from the pristine site. Denitrification were carbon limited and glucose amendment increased the denitrification activity about a 2-fold for all strains. The strain specific differences in denitrification rates increased to a 2.5-fold after carbon addition indicating that the differences in reduction rates cannot be explained by different carbon utilisation rates but rather reflect innate differences in the reductases of the strains. A preliminary identification of the molecular target for adaptation was performed with artificial electron donors and electron acceptors for all enzymatic steps in the denitrification pathway. Nitrous oxide reductase activity was significantly higher in denitrifiers from the nitrate contaminated sites. This suggests that nos genes may be the molecular target, possibly by mutation or gene duplication for adaptation to high nitrate concentrations. Two anaerobic denihifiers from each of the contaminated sites were capable of aerobic denitrification indicating that high nitrate concentrations may select for strains that denitrifies in the presence of both oxygen and nitrate. Microcosm experiments with fertilised coniferous forest soil showed that the dominating fate of added "NOj and 'wwas immobilisation in microorganisms or organic matter. The activity of denitrifiers, nitrate ammonifiers and nitrifies were negligible. suggesting that microorganisms that immobilises or mobilises nitrogen control the major fate of fertiliser nitrogen. Kq worb Denitrification, dissimilatory nitrate reduction to ammonium, nitrification, bacteria, adaptation, selection, nitrate, nitrite, nitric oxide, nitrous oxide, reductase. groundwater, aquifer, forest soil, microcosm, fertiliser, '%NO;, '%EL+ aanifiitionsyst~arld/ormdextmnr(iiany) T Supplanmtay bibtiognphical information I weEnglish

ISSN and key title I ISBN 91-7105-088-4 Recipient's notu

~~ D&iution by (meand addrrss) ' Christer Bergwall, Department of Ecology, Chemical Ecology & Ecotoxicology, S-223 62 Lund, Sweden of tfieabmaa of theaboborr-mentioneddinatation,hsebygrantttolnrcfcrrnce e the abstract of the rbore-mentioneddinatstion September 30,1997 Date Denitrification as an adaptive trait in soil and groundwater bacteria

Christer Bergwall

FK,W

Akademisk avhandling, som for avlaggande av filosofie doktorsexamen vid matematisk- naturvetenskapligafakulteten vid Lunds Universitet, kommer att offentligen forsvaras i Bli Hallen, Ekologihuset, Solvegatan 37, Lund, fredagen den 14 november 1997 kl. IO”.

Fakultetens opponent: Dr. Jan Sorensen, Institut for 0kologi og Molekylaer Biologi, Kgl. Veterinaer- og Landboh~jeskole,Frederiksberg, Kobenhavn, Danmark. Avhandlingen kommer att fdrsvaras pH engelska. Denitrification as an adaptive trait in soil and groundwater bacteria

This thesis is based on the following papers which are referred to by their Roman numerals.

I. Bengtsson G. and Bergwall C. (1995). Heterotrophic denitrification potential as an adaptive response in groundwater bacteria. FEMS Microbiol. Ecol. 16307-3 18.

11. Bergwall C. and Bengtsson G. Phenotypic plasticity in groundwater denitrifiers. (submitted)

111. Bergwall C. Anaerobic N20-reductase activity and aerobic denitrification distinguish denitrifiers from nitrate contaminated and pristine aquifers. (submitted)

IV. Bengtsson G. and Bergwall C. The fate of "N labelled nitrate and ammonium in a fertilised forest soil. (submitted)

Paper I is reprinted with permission from the publishers.

Akademisk avhandling, som fir avlaggande av filosofie doktorsexamen vid matematisk- naturvetenskapliga fakulteten vid Lunds Universitet. kommer att offentligen forsvaras 5 B11 Hallen, Ekologihuset, Solvegatan 37, Lund, fredagen den 14 november 1997 kl.

Fakultetens opponent: Prof. Jan Sgrensen, Institut for Okologi og Molekylrer Biologi, Kgl. Veterinrer- og Landboh~jjeskole,Frederiksberg, KEjbenhavn, Danmark. Avhandlingen kommer att forsvaras pi engelska. Contents Page

Introduction Aim of thesis 7 Environmental concern 7 Nitrogen transformation . Processes 8 Nitrate reducing microorganisms 10 Aerobic denitrification 10 Enzymes associated with denitrification 12 Ecological aspects of adaptation to nitrate 14 Molecular mechanisms for genetic adaptation Vertical and horizontal processes 15 Methodology for adaptation studies Denitrification measurements 15 Isolation of denitrifiers 16 Acetylene inhibition 16 Adaptation to environmental stress and nitrate Environmental stress 17 Nitrate 17 Possible mechanisms for adaptation to nitrate 18 Phenotypic plasticity of nitrogen transformers 20 Biological denitrification as a tool for remediation Heterotrophic denitrification 21 Autotrophic denitrification 22 Biorernediation of polluted aquifers with nitrate as electron acceptor 22 Biorernediation and microbial release 23 Fate of nitrate in contaminated aquifers and fertilised forest soils Groundwater versus forest soil 24 26 C/N ratio determining the fate of nitrate Future perspectives and research 27 References 28 Acknowledgement mainly in Swedish 34

Introduction

Aim of thesis

The focus of my thesis is on selection and adaptation processes in bacteria with emphasis on denitrifying bacteria in soil and groundwater. Other nitrogen transformation processes such as dissimilatory nitrate reduction to ammonium and nitrification in forest soil will be briefly discussed. Adaptation of denitrifiers to different environmental factors such as temperature, pH and salinity have been studied to some extent and will be discussed in more detail below. Concerning selection processes in denitrifiers the current view is that denitrifying strains are in general selected for competitive carbon utilisation rather than for the denitrifying property (see Tiedje 1988 for detailed discussion). The ability to compete for organic carbon under aerobic conditions is considered the main factor that regulates denitrifier distribution and numbers in most environments. However, in certain environments with high nitrate concentrations such as groundwater aquifers in agricultural areas, forest soils that have been fertilised with nitrogen fertiliser to improve plant growth or surface waters and groundwaters receiving treated sewage, a selection based on nitrate rather than carbon might be possible. The nitrate in contaminated environments may act as a driving force for selection of denitrifiers to increase their survival and fitness by utilising nitrate as terminal electron acceptor in anaerobic respiration. The few studies performed related to adaptation to nitrate will be discussed below as well as possible mechanisms for adaptation in groundwater denitrifiers. Furthermore, the potential of microbial release to remediate nitrate contaminated aquifers is an interesting applied aspect and the aim is to use adapted denitrifying strains as bioremediators or as an complement to other treatments. The advantages and disadvantages of in situ techniques and the potential use of microbial release will be explored in detail later in this thesis.

Environmental concern

Nitrate has become a pollutant in many environments e.g. groundwater and surface waters, mainly due to anthropogenic activities such as intensive fertilisation of agricultural soils and release of treated sewage to surface waters. Groundwater is an important source of drinking water in most parts of the world and nitrate contamination of water is likely to affect peo- ple in the future. Nitrate contamination is associated both with point sources and regional fertilisation and land irrigation practices (Power and Schepers 1989, Spalding and Exner 1993). Drinking water with nitrate concentrations as high as 40 mg NO<-N I-' is a health hazard, es- pecially for infants. The low acidity of the gastric juice in infants is favourable for the reduction of nitrate to nitrite. Nitrite enters the blood and oxidises iron in the haemoglobin and causes methemoglobinemia, also called blue baby (Bruning-Fann and Kaneene 1993a.b). The haemoglobin is unable to bind oxygen and symptoms of internal suffocation arise. Nitrite can also react with secondary amines to form carcinogenic nitrosamines (Bruning-Fann and Kaneene 1993a,b). Another concern is the role of nitrous oxide contributing to the global greenhouse effect and to the destruction of the protective ozone layer in the stratosphere (Robertson 1991, Lammel and Grass1 1995). Finally, subsurface discharge of nitrate from groundwaters to surface waters is mentioned, since nitrate influx to lakes and coastal marine sediments is known to occur (Johannes 1980, Capone and Bautista 1985. Staver and Brins- field 1996). High nitrate concentrations in lake and sea water have been identified as one of the major factors influencing eutrophication.

7 Organic N

It 4--- fixationNitrogec NH4' \ \ 1I ;- Assimilatory NO, Denitrification I reduction N2+ N20 t NO + NO; ...... I- Nitrification Dissimilatory NO,- J --- reduction to NH,+ Abiotic , ..e. / processes d.... , NO,- __--.. .

Fig. 1. The inorganic nitrogen cycle is composed of several linked biological processes and one abiotic process. Assimilatory and dissimilatory nitrate reduction to ammonium, nitrification, and nitrogen fixation are nitrogen conserving processes whereas denitrification is the only process that returns fixed nitrogen to the atmosphere. In the abiotic process, nitrogen and oxygen, catalysed by lightning or by internal combustion in engines, form NO, which end up as nitrate in the terrestrial environment.

Nitrogen transformation

Processes

Most bacteria are aerobic and use oxygen as the terminal electron acceptor in the electron transport chain, reducing it to water. However, there are facultative and obligate anaerobes in some environments that possess specialised electron transport chains which enable them to utilise various inorganic and organic compounds as alternatives to oxygen. These include nitrogen and sulfur compounds, ferric iron, and fumarate. Nitrogen is cycled through plants, animals and microorganisms in a complex series of biological and chemical processes, known as the nitrogen cycle (Fig. 1) In this cycle, denitrification is the only pathway that returns fixed nitrogen to the atmosphere, thus completing the nitrogen cycle. Reduction of nitrate in anoxic environments is dominated by two dissimilatory processes: respiratory denitrification, which refers to the microbially mediated process whereby nitrate is reduced to nitrogen gas and released to the atmosphere and dissimilatory nitrate reduction to ammonia (DNRA) in which nitrate is reduced to ammonium and excreted to the environment (Knowles 1982, Tiedje 1988, Cole 1990). Assimilatory nitrate reduction in which nitrate is reduced to ammonium for use as a nitrogen source for growth, can also occur in anoxic environments, but is considered to be insignificant since the ammonium concentration often is high in anoxic environments due to

8 lack of nitrification (Tiedje et al. 1981, Tiedje 1988). In the respiratory process, nitrate serves as the terminal electron acceptor in the electron transport chain, resulting in generation of ATP through electron transport phosphorylation. The main characteristics of the processes are summarised in Table 1. Nitrification is a nitrogen conserving process which is performed by both autotrophic and heterotrophic bacteria (Jetten et al. 1997). Nitrification is an aerobic process and occurs read- I ily in well-drained soils at neutral pH (Prosser and Cox 1982). It is inhibited by anaerobic ' I' conditions and in highly acidic soils (Prosser and Cox 1982). In autotrophic nitrification, conversion of ammonium to nitrate occurs in a two step process as a sequential action by two separate groups of bacteria, the ammonium oxidisers and nitrite oxidisers. Ammonia and nitrite are used as electron donors and oxygen as electron acceptor. Recent work have demonstrated that some autotrophic nitrifiers also denitrifies under oxygen limiting growth conditions. Nirrosornonas europea and a Nirrosornonas sp., which are aerobic ammonium oxidising

4- nitrifiers, oxidised ammonium to nitrite aerobically coupled to nitrite reduction to nitric oxide, nitrous oxide and nitrogen gas under oxygen limiting conditions (Poth and Focht 1985, Poth 1986, Remde and Conrad 1990). Nitrobacter strains, which are aerobic nitrite oxidisers, have been shown to grow aerobically by nitrite oxidation and anaerobically by denitrification (Freitag et al. 1987). Bock et a]. 1995 reported that N. europaea and Nirrosornonus ercrrophu were able to nitrify and denitrify simultaneously under oxygen limiting growth conditions with significant nitrogen losses. Heterotrophic nitrification has been observed in bacteria, fungi and algae producing nitrite or nitrate without gaining energy from the process (Kuenen and Robertson 1994). An organic carbon source is generally needed to complete the reactions. Many heterotrophic denitrifiers tend to couple their activity to heterotrophic nitrification (Castignetti and Hollocher 1984). The bacteria capable of both processes were Pseudomonas and Alcaligenes species, which are common in most environments. It was concluded that nitrifier denitrification is advantageous to bacteria under fluctuating environmental conditions. These coupled reactions are an example of the metabolic flexibility of nitrogen transformers under transition conditions. In oxic environment heterotrophic or autotrophic

Table 1. Biological nitrate reduction pathways in microorganisms. Modified from Tiedje (1988). Process Metabolic pathway Mainly Organisms possessing regulated process by:

~~ ~~ Dissimilatory pathways: Denitrification NO; + NO; 3 NO + N20 + N2 02 Aerobic bacteria capable of anaerobic growth, certain fungi

Dissimilatory NO; 4 NO; + NHJ* 02 Anaerobic and faculta- nitrate reduction to tively anaerobic bacteria ammonium Assimilatory Patltway: Assimilatory NO; 4 NO; + NH; NH;, Bacteria, fungi, algae, . I, nitrate reduction to organic N plants , ammonium

9 nitrification generates nitrite or nitrate which then will be available for denitrification in the transition phase from oxic to anoxic growth. For a more detailed discussion about different aspects of nitrification and nitrifier denitrification the reviews of Kuenen and Robertson 1994, Hooper et al. 1997 and Jetten et al. 1997 are recommended.

Nitrate reducing microorganisms

Bacteria capable of nitrate reduction are widely distributed in nature, including habitats such as soil, fresh water, groundwater, marine waters, sediments, waste treatment systems and the human and animal gastrointestinal tract (Knowles 1982, Tiedje 1988). The energy source of nitrate reducers include all three classes known to be used by microorganisms: organic (organotrophs), inorganic (litotrophs) and light (fototrophs) (Tiedje 1988, Beauchamp et al. 1989). Most of them are organotrophs and use a wide variety of organic acids, carbohydrates and other organic compounds as carbon and energy source. The denitrification capacity is spread among a wide variety of physiological and taxonomical groups. The most common denitrifies in nature are gram-negative species of Pseudoinonas followed by the closely related Alcaligenes. Gram-positive bacteria such as the thermophilic Bacillus stearothermophilus are also capable of denitrification (Ho et al. 1993). Denitrification is not exclusively a bacterial process since fungal denitrification has been observed in Fusarium oxysporum and related fungi (Shoun et al. 1992). Earthworms collected from a beech forest and an acidic oak-beech forest soil have recently been shown to carry denitrifiers in the gastrointestinal tract (Karsten and Drake 1997). The average culturable denitrifier numbers in gut material from Lumbricus rubellus and Octalasium lacteum were 7 x IO’ and 9 x IO6, respectively, which was 256 and 35 times higher than the culturable number in the surrounding soil. Nitrous oxide was produced under both oxic and anoxic conditions by gut material and living earthworms. Acety- lene addition increased nitrous oxide production suggesting that denitrifiers were responsible for nitrate reduction. Nitrous oxide emission rates from the earthworms were calculated to ac- count for about 16 and 0.25%. respectively, of the total production in these soils suggesting that earthworms are ”mobile hot spots” for denitrification in terrestrial environments. Micro- organisms capable of DNRA have been found in anoxic environments such as anoxic sediments, animal gastrointestinal tracts and sludge digestors (Tiedje 1988, Cole 1990). In con- trast to respiratory denitrification, DNRA is predominately performed by fermentative bacteria, such as Clostridia, and rumen bacteria such as Wolinella and Selenomonas species as well as a number of facultatively anaerobic species belonging to the Enterobacteriace (Tiedje 1988, Cole 1990). The first step in DNRA, nitrate reduction to nitrite, is coupled to energy production by electron transport phosphorylation or by the generation of an electrochemical proton gradient (Cole 1990). The purpose of the conversion of nitrite to ammonium is unclear and several explanations have been proposed (see discussion in Tiedje 1988, Cole 1990).

Aerobic denitrification

Denitrification has been considered to be exclusively an anaerobic process since expression and activity of denitrification enzymes are repressed completely under oxic growth conditions (Wu et al. 1994). However, recent studies have shown that a number of anaerobic denitrifiers are capable of aerobic denitrification. Robertson and Kuenen (1983) reported that the gram-negative Thospaera pantotropha isolated from a desulfurizing, denitrifying waste-water treatment system, expressed denitrifying enzymes in the presence of oxygen concentrations up

IO to 90% of air saturation. Lloyd et al. 1987 studied the effect of oxygen on denitrification in eight different anaerobic denitrifiers. Three of them were Pseudomonas species isolated from a marine sediment. All bacteria could denitrify in presence of oxygen that approached or ex- ceeded air saturation values. The biochemical basis for the aerobic denitrification in Pseudo- monas is not known but, two distinct nitrate reductases located in the periplasmic space and in the cytoplasmic membrane were shown by physiological means in the gram-negative Thio- spkera panrotropha (Bell et al. 1990). Isolation of a novel type of denitrification enzyme, a nitrate reductase (Nap), located in the periplasmic space of T. panfofrophaand active under oxic growth conditions was reported by Berks et al. (1994). Nap of T. panfofrophais not involved in anaerobic denitrification since expression of the periplasmic nitrate reductase is repressed during anoxic growth (Bell et al. 1990). Furthermore, it is not involved in assimilatory nitrate reduction since expression of Nap is not repressed by ammonium (Berks et al. 1995). Evidence for a possible involvement of Nap in aerobic denitrification comes from the simple fact that nitrate reduction in the periplasm is not inhibited by oxygen (Berks et al. 1995). Nap has further two properties that separates it from the membrane-bound nitrate reductase. Nap is not inhibited by azide and it can not reduce chlorate as an alternative substrate (Bell et al. 1990, Berks et al. 1995). Carter et al. (1995) found 29 strains, isolated from soil and freshwa- ter sediments, that were capable of aerobic nitrate reduction. All strains expressed a nitrate reductase located in the periplasmic compartment, indicating the presence of Nap. The numbers of culturable bacteria capable of aerobic nitrate reduction ranged from IO4 to IO' g" soil or sediment and were equally or more abundant than culturable anaerobic denitrifiers. This suggests that aerobic nitrate reduction and, possibly, aerobic denitrification is not limited to a few specialised strains but rather commonplace in different environments. In paper III, two strains from each of the two nitrate contaminated groundwater aquifers were capable aerobic nitrate, nitrite, nitric oxide and nitrous oxide reduction indicating that they are aerobic denitrifiers. Aerobic nitrate reductase was located in the periplasm indicating the presence of Nap. No aerobically active isoenzymes of Nir, Nor and Nos could be identified and their individual location suggests that they also are involved in anaerobic denitrification as suggested for T. partorropha (Bell and Ferguson 1991, Moir et al. 1993, Berks et al. 1993). No aerobic activity was found in denitrifiers isolated from the uncontaminated site. An alternative explanation for the function of Nap in aerobic nitrate reducers was suggested by Berks et al. (1995) and Zumft and Korner (1997). The presence of two types of nitrate reductases may be advantageous to the strains since electrons generated in respiration can be transported to two potential terminal electron acceptors. As the transport of nitrate across t. ,o , the cytoplasmic membrane is inhibited by oxygen, the periplasmic nitrate reductase may serve as a bridge in the transition phase from oxic to anoxic growth conditions acting as an electron sink to maintain cellular redox balance in the transition phase. This has been implicated as a most likely explanation for the presence of Nap in the anaerobic denitrifier Alcaligenes ercrro- phis (Siddiqui et al. 1993). Several bacterial species are capable of simultaneously utilising oxygen and nitrate as terminal electron acceptors. (Robertson and Kuenen 1984, Bonin and Gilewicz 1991. Patureau et al. 1994). Using both oxygen and nitrate, Comnrnonas sp. SGLY2 grew about ten times faster compared to growth with only nitrate (Patureau et al. 1994). In paper III, simple growth experiments in a minimal medium with and without nitrate is performed to test the hypothesis that the simultaneous use of oxygen and nitrate increased the fitness of aerobic denitrifiers. The lag phase and exponential growth phase was almost identical between the strains regardless of the presence or absence of nitrate. Strains growing with both oxygen and nitrate reached stationary phase earlier with lower growth yields, which suggests that growth in the presence of both oxygen and nitrate was limited by trace minerals needed for synthesis of Table 2. Reactions of reductases that are active in the denitrification and reduction potentials (Eo) for each redox pair. Modified from Averill (1996). Enzyme Enzymatic reaction Redox pairs Eo'

~~~ (VI Nar NOi + 2e' + 2H' + NO; + H20 NOi/NO; +0.42 Nir NO; +e- + 2H'+ NO + H20 N0;MO +0.37 Nor 2N0 + 2e- + 2H' + N20 + H20 2NOM20 +1.18

Nos N20 + 2e' + 2H* --f N2 + H20 NzON +1.77 enzymes active both under oxic and anoxic growth conditions. The advantage of utilising two electron acceptors was demonstrated in competitive growth experiments between T. pantotropha, which denitrified under both oxic and anoxic conditions and Paracoccus denitrificans, which denitrified only under anoxic conditions (Robertson and Kuenen 1992).

Enzymes associated with denitrification

The enzymology of denitrification will only be summarised here and detailed information on different aspects of the enzymes of denitrification can be found in the reviews of Hochstein and Tomlinson (1988), Stouthamer (1988), Ferguson (l994), Berks et al. (1995), Averill (1996) and Zumft and Korner (1997). The reduction of nitrate to nitrogen gas proceeds in four steps and is catalysed by specific enzymes designated reductases. Denitrification has long been considered a strict anaerobic process since oxygen tends to inhibit the transport of nitrate over the cytoplasmic membrane (Noji and Taniguchi 1987, Hernandez and Rowe 1988, Wu and Knowles 1994) and repress the expression and activity of denitrification enzymes (Wu et al. 1994). Anaerbiosis alone induces reductase expression in some strains but not in others (Korner and Zumft 1989, Coyne and Tiedje 1990b). Nitrate induces the expression of nitrate reductase but also other reductases and in some cases all (Korner and Zumft 1989). Nitrite and nitrous oxide stimulate mainly the expression of nitrite and nitrous oxide reductase, respectively (Arai et al. 1991, Korner and Zumft 1989, Coyne and Tiedje 1990b). The reactions of the reductases and the reduction potentials are summarised in Table 2. Judging from the reduction potentials, nitrate, nitrite, nitric oxide and nitrous oxide reduction coupled to oxidation of electron donors provide a sufficiently strong driving force to power the electron transport chain in bacteria. The organisation of reductases in gram-negative denitrifiers is shown in Figure 2. The nitrate reductase (Nar) is a membrane-bound enzyme which reduces nitrate to nitrite. The active site of the enzyme is located at the cytoplasmic side of the membrane which means that nitrate has to be transported across both the periplasmic and the cytoplasmic membranes. The enzyme consists of three subunits, a,p, and y. Subunit a is believed to contain the active site and has been proposed to function as transporter of oxygen when nitrate is reduced to nitrite. Subunit p contains iron-sulfur-clusters, which potentially serve as electron mediators between the subunits. Subunit y is a cytochrome b and is proposed to function as a membrane anchor for the nitrate reductase complex. The periplasmic nitrate reductase active

12 outside NO,- NO,- NO N*O N, I+ I+ t $1 +I fl

membrane

cytoplasm

Fig. 2. The cellular arrangement of denitrification enzymes in most gram-negative bacteria. Nitrate reductase (Nar) resides in the cytoplasmic membrane with its active located at the cytoplasmic side, necessitating transport of nitrate across both the periplasmic and the cytoplasmic membranes. Nitrite (Nir) and nitrous oxide (Nos) reductase are soluble enzymes located in the periplasmic space. Nitric oxide reductase (Nor) are a membrane-bound enzyme. All three have active sites accessed from the periplasmic space.

in aerobic denitrification consists of two subunits, designated Nap A and B (Berks et al. 1995, Zumft and Korner 1997). Nap A is proposed to contain the active site of nitrate reduction. The function of Nap B is unknown but loss of the Nap B subunit leads to enzyme inactivation (Berks et al. 1995). Nitrite is transported back into the periplasmic space where nitrite reductase (Nir) reduces it to nitric oxide, which is the first gaseous compound released during denitrification. Two distinct types of Nir have been found in denitrifiers. One is a two heme containing enzyme, the cytochrome cd~,and the other is a coppercontaining enzyme. The C-type $ heme is believed to function in electron transfer from electron donors to the dl heme. The dl ' I' heme is assumed to contain the active site of nitrite reduction. The copper type of Nir contains ,. . three identical subunits and two copper atoms per subunit termed, type I and type II copper. Type I copper is assumed to serve as electron mediator from electron donor to type II copper, analogous to that of C-type heme in the cytochrome cdl nitrite reductase. Type II copper is located in the active site, which binds and reduces nitrite. Nitrite reductases of the cdl-type have been found to be most common in environmental isolates of numerically dominant denitrifiers from soil, sediment and sewage. However, the copper nitrite reductase seems to be widely distributed among bacteria in different environments (Coyne et al. 1989, Coyne and Tiedje 199Oa). No bacteria have been reported to contain both types and it appears that there is no relationship between species and what kind of nitrite reductase is present (Berks et al. 1995, Averill 1996). Nitric oxide is further reduced by the membrane-bound nitric oxide reductase (Nor) with the active site at the periplasmic side of the membrane. Nor is a two subunit protein that contains a cytochrome c and b. The reduction of nitric oxide is the least well characterised of the enzymatic steps, partly because of uncertainties about the chemical reactivity of

13 nitric oxide as well as doubts concerning its role as an intermediate in the denitrification. It has now been verified that Nor is a part of the denitrification pathway (Ye et al. 1994 and references therein). Nir and Nor reductase activity have been suggested to be coupled reactions and function as a multienzyme complex in which produced nitric oxide is proposed to be directly channelled to the active site of the Nor. Nitric oxide is a toxic intermediate which easily reacts with heme and non-heme iron-containing proteins. (Ye et al. 1994, Averill 1996). Nitric oxide is returned into the periplasm where nitrous oxide reductase (Nos) reduces it to nitrogen gas which is released to the atmosphere. Nos is a two subunit enzyme with four copper atoms per subunit. The Cu atoms are organised into two groups, designated CUAand CUZ. The CuA is considered to function in electron transfer from electron donors to the active site at Cuz.

Ecological aspects of adaptation to nitrate

Evolution is based on changes in the relative frequency of genotypes in a population over a period of time. Natural selection, which is considered one of the driving processes leading to evolution of organisms that are adapted to their environment, refers to the environmentally directed, stabilising or disruptive change in the populations genetic structure. Adaptation is defined as the evolutionary process by which an organism increase its fitness under the prevailing environmental conditions, and as the specific genetically determined trait responsible for the fitness increase (McNaughton and Wolf 1979). In molecular terms, adaptation refers to the evolution of metabolic capabilities produced by genetic changes of indigenous genes, or recombination of foreign genes, or mutations producing new phenotypes which are selected for in the environment. Adaptation is often expressed in response to changes in the environment and only those modifications that positively affect survival and fitness are designated adaptive. The response of denitrifiers to increased nitrate concentrations include various mechanisms, primarily, behavioural and physiological. A behavioural as well as physiological response to nitrate is positive chemotaxis expressed by denitrifiers reported by Kennedy and Lawless (1985). The first observable chemotactic response was at about IO5 M for both nitrate and nitrite and a maximum response was found at M. The taxis response was exhib- ited both under oxic and anoxic growth conditions and independent of previous growth on nitrate. The authors tested chemotaxis in indigenous soil bacteria using M nitrate as the at- tractant. Soil bacteria from all soils, predominantly Pseudomonas sp, showed a significant response to nitrate. It was concluded that the chemotactic response may be one mechanism which enables denitrifiers to successfully compete for available nitrate and nitrite. Intensive fertilisation of agricultural soils in southern Sweden have been performed during the last 20-30 years and has resulted in accumulation of nitrate in aquifers. Selection of denitrifiers in environments with high nitrate concentrations may occur since nitrate may act as a selective agent to alter gene frequencies in natural denitrifier populations. Selection may be accompanied by a community shift with a larger part of the community being denitrifiers with enhanced activity. In essence, a denitrifier that can utilise increased nitrate concentrations more efficiently than its neighbours, becomes more competitive and can allocate increasing amounts of energy to growth and reproduction which would result in an increase in fitness. The community shift may be reflected in the isolation frequencies in paper I in which about 50% of strains isolated from the nitrate contaminated aquifers were denitrifiers compared to about 20% for the pristine site. Another adaptive response of denitrifiers in nitrate contaminated aquifers may be the ability to denitrify in the presence of oxygen. The nitrate concentration of the contaminated sites was 5-7 times higher than the oxygen concentration. Oxidation of organic carbon using

14 either oxygen or nitrate yield comparable AGO values, -18.76 and -17.26 kcal electron", respectively (Korom 1992). which indicates that reduction of high nitrate concentrations may be energetically rewarding for the strains. It is possible that the strains which denitrifies in the presence of oxygen may have a selective advantage over the purely anaerobic denitrifiers ac- 'I companied by a community shift towards aerobic denitrifiers analogous to the shift towards anaerobic denitrifiers with high denitrifying activity.

Molecular mechanisms for genetic adaptation

Vertical and horizontal processes

A distinction between vertical and horizontal evolutionary processes is necessary when '. genetic mechanisms involved in evolution of microorganisms are considered. Vertical processes are based on random mutations which are spread in populations when cells divide. Mu- tations of the DNA are mostly harmful or neutral, although occasionally beneficial changes occur followed by selection for the resulting phenotype. Mutations are spontaneous events, i.e. they happen randomly with no way of knowing when or in which cell a mutation will occur. Mutations are also random in the sense that the occurrence is not related to any adaptive advantage it may confer to the organism in its environment. Two important mechanisms for mutational alteration of the DNA are errors occurring during replication (transcription) and spontaneous or induced alteration of one or more base pairs of the DNA. The chemical and physical properties of each protein are determined by its amino acid sequence, so a single amino acid substitution, a point mutation, is capable of altering the structure of a protein. If the change occurs at a critical point, the protein may be inactivated or reduced in activity. However, the change may also enhance the properties of the enzyme, resulting in a more efficient utilisation of the substrate or even gain new metabolic capabilities that may be essential for survival. Horizontal processes is based on exchange of DNA between two different cells or between chromosomes and extrachromosomal complexes such as plasmids. Bacteria are haploid organisms and do not depend on sexual reproduction to multiply and adapt to changing environments. The basis for adaptation by acquiring new genetic components is therefore different compared to that in eukaryotic organisms. Genetic material can be exchanged in several ways: (I) By transformation, in which uptake of free DNA from the environment can be incorporated by recombination. (2) Conjugation, in which a donor strain link up (direct contact) with a recipient strain. DNA is transferred from the donor by use of sex pili to the recipient. The DNA can then be incorporated by recombination. This is also a common way for plasmids to spread within and between bacterial species. (3) By transduction, in which bacterial viruses can function as transfer vectors for DNA from a donor strain.

Methodology for adaptation studies

Denitrification measurements

The approach taken to study in situ denitrification activity in general is similar for most experimental work. Intact soil cores or sediment slurries are used to quantify the activity. This also applies for the few investigations of adaptation to nitrate (King and Nedwell 1987, Smith and Duff 1988). Such experiments can be considered to address whole community denitrification and indicate difference in activity between field sites. The drawback with that approach is the uncertainty whether a few strains or all strains are the source of activity. No insight is gained about the variation in activity among the members of the community. Ultimately, if the aim of a study is to demonstrate the potential adaptation to a compound, the use of soil cores or sediment slurries is impractical. Adaptation can only be studied when studying individuals and not whole communities. Experiments like those described above are valuable in the sense that they give a preliminary assessment of the differences in denitrification between sites, but nothing more. The approach chosen in my work was to use single strains to elucidate the pat- terns of adaptation to nitrate. The drawback of my studies has been the difficulty to find and compare the performance of the same bacterial species or strain in all of the field sites. This was an impractical task, so I used several strains to give a distribution of denitrifying activity within and between nitrate contaminated and pristine sites. A kinetic approach was used to quantify the denitrifying activities and the obtained rates were used to compare the activity of strains or soils from different sites. The kinetic properties of single strains were used to express adaptation to in situ N concentration of their habitat.

Isolation of denitrifiers

Oxic incubation was chosen to isolate bacterial strains since denitrifiers, generally, are aerobic bacteria. Even if a proper isolation methodology was adopted it may have limited the number of culturable denitrifiers and selected for strains that were not representative for the whole community. The isolation of bacteria from groundwater sediments is influenced by the type of growth media that is used. Hirsch and Rades-Rohkohl (1988) isolated groundwater bacteria using two different media, PM and PYGV, which resulted in similar numbers of viable cell, but affected the diversity of microorganisms. PM selected for spirilla whereas no spirilla were detected using PYGV. The frequency of rods was about 90% using PYGV and about 60% on PM. The determination of the number of strains using culture media is in general inaccurate for at least three reasons: (1) Isolation on agar media is selective for different metabolically active strains. (2) The assumption of one cell per colony is violated by the presence of aggregated cells. (3) Some viable cells are non-culturable resulting in variable plating efficiencies (Roszak and Colwell 1987). All these factors contribute to an underestimation of bacterial numbers in subsurface soils.

Acerylene inhibition

The acetylene inhibition technique was used in the experiments (Knowles 1990). Nitrous oxide reductase (Nos) activity, one of the enzymes active in denitrification, is inhibited by acetylene and nitrate are quantitatively converted into nitrous oxide in the presence of acetylene. This provides the basis for a reliable method to quantify denitrification since the back- ground concentrations of nitrous oxide are considerably lower than nitrogen gas concentrations. Two major problems with the technique have been identified. Acetylene inhibits ammonium monooxygenase activity in autotrophic nitrifiers (Klemedtsson et al. 1990, Knowles 1990). Inhibition of nitrifiers may affect denitrification in long-term incubations since nitrate will not be supplied to denitrifiers. Acetylene may also be degraded by soil microorganisms. Certain strains can utilise acetylene as the sole carbon and energy source, producing, e.g. al- cohols and organic acids (Knowles 1990). The released compounds can be utilised by heterotrophic denitrifiers which will result in an overestimation of the denitrification potential. In

16 addition, the decrease in acetylene concentration may relieve the inhibitory effect on Nos and parts of the reduced nitrate will accumulate as nitrogen gas. The effects are, generally, experi- enced when long-term whole community studies are performed. In my experimental work, single groundwater strains and forest soils were used in short-term incubations, normally. less than six days, which most likely minimised the problems with acetylene.

Adaptation to environmental stress and nitrate

Environmental stress

The adaptation of denitrifiers to environmental stress has been studied to some extent and includes adaptation to temperature (psycrophiles to thermophiles), high salinity (halophiles) and varying pH. Saad and Conrad (1993) isolated nitrate reducers from an Egyp- tian soil and from two German soils. The majority of nitrate reducing strains grew best at temperatures of 2530°C whereas three denitrifying strains had optimal growth and denitrification at 8, 25 and 40°C. respectively. Hollocher and Kristjinsson (1992) screened for thermophilic denitrifying bacteria in hot springs in south-western Iceland. Nitrate reducing bacteria and denitrifiers were found that grew at 7OoC in nutrient media at pH 8. Moderate and ex- treme halophilic strains capable of denitrification have been reported to tolerate salt concentrations up to about 5 M NaCl (Tomlinson et al. 1986, Denariaz et al. 1989, Shieh and Liu 1996). Parkin et ai. (1985) measured denitrification rates and enzyme activity in two agricultural soils, one of which had a pH of about 4 after fertilisation for 20 years with acidic ammonium salts, and one that had been limed resulting in a pH of about 6. The denitrification rates were of similar magnitude, and although in situ denitrifying enzyme activities were higher in the neutral soil, substantial enzyme activities were also detected in the acid soil. Denitrifica- tion rates showed distinctly different pH optima, 3.9 in the acid soil and 6.3 in the neutral soil, which suggested that the fertilised soil had selected for acid tolerant denitrifying populations.

Nitrate

Studies of potential adaptation to high nitrate concentrations are scarce and often not the objective of the study. King and Nedwell (1987) investigated denitrification activity in an estaurine sediment in the Colne River Estuary, UK. The river was contaminated with nitrate from a sewage treatment plant which had created a contaminant plume within the estuary. Sediment and water were sampled from ten sites within the plume and used in slurry experiments. Denitrifying activity increased with increasing nitrate concentration, whereas DNRA activity was low at all sites. The authors suggested that denitrifiers had been selectively modified in response to the long-term nitrate loading of the estuary. Smith and Duff (1988) investigated the denitrification activity of a nitrate contaminated sand and gravel groundwater aquifer at Cape Cod, USA. The groundwater aquifer was contaminated by treated sewage and had a contaminant plume of over 3.5 km. Denitrification rates were highest closest to the contaminant source and decreased with increasing distance. Denitrification was carbon limited and no DNRA activity was observed. The correlation between the denitrification activity and nitrate concentrations suggests that adaptation of denitrifiers to in situ nitrate concentration may occur.

17 Possible mechanisms for adaptation to nitrate

Genetic adaptation to xenobiotic compounds may give some clues on potential mechanisms for denitrifier adaptation. Adaptation of microorganisms to xenobiotic compounds has been intensively studied since bacteria and fungi are the only efficient degraders of various organic pollutants. Horizontal transfer of genetic information has been identified to play a major role in the adaptation of microorganisms to xenobiotic compounds (van der Meer et al. 1992). The occurrence of self-transmissible plasmids, coding for catabolic enzymes, in natural environments is the basis of an extensive gene pool which may be distributed among indigenous microorganisms. Vertical processes, mainly point mutations, are identified as another important process, primarily, by affecting substrate specificity of degradative enzymes or the appearance of novel degradative pathways. Duplication of degradative genes or entire operons is considered to occur frequently in bacteria (van der Meer et al. 1992). The same mechanisms may be important in denitrifier adaptation to increased nitrate concentrations in contaminated groundwaters. A complicating factor when considering adaptation is the genetic variation in populations. Denitrifiers with high activity may already have been present in groundwater aquifers before the onset of modem agricultural practices. When nitrate concentrations increased in the aquifers, strains with high denitrifying activity may have been selected for simply by their ability to out-compete strains with lower denitrifying activity. The capacity for high denitrification rates in strains would then only be based on genetic variation within the population and not driven by natural selection per se. However, if we assume that selection for genotypes with high denitrifying activity has occurred, several mechanisms can be considered: Spontaneous mutations in regulatory or structural genes resulting in modified genes that increase enzyme expression rates or expression of enzymes with higher catalytic activity, increased affinity for the substrate or increased specificity for the substrate (vertical mechanism). Mutation in structural or regulatory genes in denitrifiers is possible and been selected for by the long-term selective pressure of nitrate in agricultural aquifers or fertilised forest soils. Even if the cell division rates are slow in oligotrophic environments, generation times in oligotrophic aquatic bacteria have been reported to range from 20 to 210 hr (Roszak and Colwell 1987). 20-30 years of continuous nitrogen loading is, likely, sufficient for some form of mutation to occur and to be selected for in denitrifier populations. Regulation of the reductase expression in the presence of trace concentrations of oxygen (-0.5 mg I-' Or),appear to be an important difference between the strains from the contaminated and the uncontaminated aquifers. In paper I and II, lag phases before onset of denitrification were consistently shorter for all strains isolated from the contaminated site regardless of the nitrate concentration, suggesting that the expression of anaerobic reductases as well as the enzyme activity was tolerant to low levels of oxygen, possibly, by low level constitutive expression of one or several reductases. Constitutive expression of Nar and Nos is known for Achomobacrer cycloclastes (Coyne and Tiedje I990a) and for Nos in Pseitdomorzas sfutzeri (Korner and Zumft 1989). Exchange of reductase genes through transfer and recombination resulting in entirely new denitrification genes (horizontal mechanism). Transfer and recombination assumes that new genes are acquired from other microorganisms by direct or indirect transfer and exchange of DNA. Genes coding for denitrification enzymes may be transferred from strains within the bacterial community or may originate from strains that have immigrated from other environments such as surface waters or the vadose zone, directly above the groundwater. Horizontal transfer may be indicated by the presence of a megaplasmid in the denitrifier Alicaligenes eu- rrophus. Denitrification in this organism is plasmid dependent and the genes for a periplasmic nitrate reductase (nap) and nosZ, the structural gene of the nitrous oxide reductase, are located on the conjugative 450kb megaplasmid pHGl (Romermann and Friedrich 1985, Zumft et al. 1992, Siddiqui et al. 1993). The megaplasmid was transferred from A. eurrophiis HI6 to a herbicide-degrading A. eutrophus JMP134, which lacked the metabolic capacity conferred by the plasmid (Schneider et al. 1988). The plasmid was stable in A. eutrophus and transconjugant cells expressed all metabolic functions associated with the megaplasmid.JMP134 The structural genes for the cytochrome cdl nitrite reductase of A. eutrophiis have recently been shown to be located on the chromosome (Rees et al. 1997). Rhizobium melilori carry the nos2 gene together with genes for symbiotic nitrogen fixation (nod) on the nod megaplasmid (Chan and Wheatcroft 1993). This suggests that horizontal transfer may be an important mechanism for acquisition of novel denitrification genes as well as a potential mechanism for adaptation to high nitrate concentrations. Presence of duplicate reductase genes and the potential of separate reductases with low and high denitrifying activity. Duplication of genes occur in bacteria and, generally, the duplicate is a copy of the original gene, thereby increasing enzyme yield (Clarke 1983). Higher growth rates give a selective advantage to the cells in which the duplication has occurred. The duplicate may evolve separately from the original gene without selective constraints and accumulation of mutations may lead to inactivation of the gene or genes expressing enzymes with modified characteristics or entirely new metabolic capabilities. Duplication of nitrate reductase genes, involved in nitrate reduction to ammonium, has been suggested to occur in Es- cherichia coli (Blasco et al. 1992, Cole 1996). The two main nitrate reductases, NRA and NRZ, are membrane-bound and composed of three subunits, a, p and y. The first one is induced by nitrate during anaerobic growth, whereas the second is constitutively induced at low levels. The isoenzymes are coded by genes located in two operons, narGHJI and narZW, which are similar in size and in genetic organisation. In paper III, Nos activity was significantly higher in strains from the contaminated sites suggesting that nos genes may be the molecular site for adaptation in strains from contaminated aquifers. All strains contained Nos in the periplasmic space, which suggests that the duplicated gene may be an identical copy of the original gene. Alternatively, the mechanism introduced by duplication of genes coding for reductases with different denitrifying activity for nitrate provides the bacteria with a more flexi- ble and energy saving system for utilisation of nitrate. The cell may express the appropriate system in response to the external nitrate concentrations. The potential molecular mechanism for this kind of duplication is, to my knowledge, unknown. The duplication event may have lead to two identical copies, but with time the duplicate could have undergone mutations, resulting in genes coding for denitrification enzymes with different properties. Appearance of aerobic denitrification and co-respiration of nitrate and oxygen. The appearance of aerobic denitrification may be an adaptive mechanism of denitrifies in nitrate contaminated aquifers. The bacteria may evolve towards co-respiration of oxygen and nitrate either by isoenzymes that are active in oxic environments or by enzymes with a broad functionality, i.e. enzymes active in the presence of both oxygen and nitrate. The presence of Nap in the groundwater denitrifiers and other strains supports the idea of separate reductases but no isoenzymes of Nir, Nor and Nos have been found so far (Bell and Ferguson 1991, Moir et al. 1993. Berks et al. 1993, paper III). The notion of broad functionality is supported by the fact that only Nap was active under oxic growth conditions and by the individual location of the other aerobically active reductases. It appears that the expression and activity of Nir, Nor and Nos in aerobic denitrifies is not sensitive to oxygen and reasonable assumption is that the enzymes are active under both oxic and anoxic growth conditions. Presence of low- and high-affinity dissimilatory nitrate transport systems. Nitrate transport across the cytoplasmic membrane could be one factor that influences denitrifying activity

19 of the groundwater strains. It is possible that strains from nitrate contaminated aquifers may have transport proteins with high activity or two transport systems with separate affinities for nitrate. Increased transport rates would effectively supply Nar with substrate. A low- and high- affinity assimilatory transport system for nitrate has been found in Klebsiella pneumoniae (Thayer and Huffaker 1982) and in the unicellular algae Cyanidium caldariurn (Fuggi et al. 1984) suggesting that assimilative nitrate transport is composed of two specific permeases active under different environmental conditions and with separate optimum activity. Several mechanisms for respiratory nitrate transport have been presented but no clear evidence for a specific transport mechanism exist to date (see discussion in Berks et al. 1995). So far, denitrifiers are not known to have a low- and high-affinity transport system for dissimilatory nitrate reduction. The advantages with a dual transport system are much the same as those described for duplicate denitrification genes coding for reductases with different activity. Uptake of nitrate can be regulated in response to the external nitrate concentration to maximise substrate supply.

Phenotypic plasticity of nitrogen transformers

Differences in the individual ability to respond to fluctuating conditions may be induced by environmental changes, Le. a single genotype may produce different phenotypes under different environmental conditions. Phenotypic plasticity is defined as plasticity genes, i.e. regulatory genes, that exert environmentally dependent control of the response of a genotype, Le. expression of structural genes (Schlichting and Pigliucci 1993, Pigliucci 1996). In general not all genes of an organisms are expressed in a certain environment. Expression of structural genes is controlled by regulatory genes that are triggered by environment specific conditions. Phenotypic plasticity in nitrogen transformers is indicated by multiple enzyme systems expressed separately under different environmental conditions. Anaerobic respiration has been extensively studied in Escherichia coli and is an excellent example of phenotypic plasticity in nitrogen metabolism. E. coli uses two separate enzymatic pathways independent of each other to reduce nitrate to ammonium. A membrane-bound nitrate reductase operates in conjunction with a cytoplasmic nitrite reductase when nitrate is abundant whereas a periplasmic nitrate reductase is active together with a periplasmic nitrite reductase when nitrite is abundant and nitrate is scarce (Cole 1996). Assimilative nitrate uptake in Klebsiella pneumoniae and in the unicellular alga Cyanidium caldariurn occur through two specific uptake systems with low and high affinity for nitrate (Thayer and Huffaker 1981, Fuggi et al. 1984). The two systems are regulated by the external nitrate concentration, the low affinity system is active in the presence of mM concentrations of nitrate and the high affinity in the presence of pM concentrations. In paper II, denitrifying strains from the pristine site had a narrow optimum of denitrifying activity at low nitrate concentrations (3-6 mg N0i-N I-') with low variation in activity between strains as well as an optimum population growth around the in situ concentration, which suggests that they have undergone stabilising selection to the continuously low nitrate concentrations of that aquifer. Denitrifying strains from the contaminated site maintained a high denitrifying activity across almost all nitrate concentrations (3-24 mg NO<-N I-') with larger variation in activity between the strains. This difference may be attributed to the pulses of high NO< concentrations originating from periodic fertilisation of the agricultural aquifers and which prevent stabilising selection. The variation in denitrifying activity between the strains from the contaminated site was not density dependent since the variation was maintained when calculated on cellular basis. Two of the five anaerobic denitrifiers used in this

20 experiment have been shown to be capable of aerobic denitrification (paper m) which may be interpreted as a plastic response in denitrifiers from the nitrate contaminated sites. The denitrifiers expressed Nap under oxic growth conditions and Nar under anoxic conditions, which clearly shows that oxygen is the main regulatory mechanism. Nap in T. panrofropha and A. eufrophus does not require nitrate for induction whereas Nar is, generally, an inducible enzyme with some exceptions (Calder et al. 1980, Korner and Zumft 1989, Coyne and Tiedje I990a, Richardson and Ferguson 1992, Warneke-Eberz and Friedrich 1993).

Biological denitrification as a tool for remediation

Heterotrophic denitrification

Nitrate has become a pollutant in many environments and nitrate concentrations in aquifers in areas with sandy soils used for arable crops, cattle grazing, and irrigated cropping of vegetables can exceed the maximum admissible drinking water limit of I 1.3 mg N0i-N I-' set by the European Community (Anonymous 1980) and the maximum contaminant level of IO mg NO;-N I" established by the U.S. Environmental Protection Agency (Fried 1991, Spald- ing and Exner 1993). The options for remediating groundwater includes physicalkhemical processes such as ion exchange and reverse osmosis. The major drawback of chemical processes are the lack of selectivity for nitrate. All ionic forms are replaced leaving a low quality drinking water that has to be treated for human consumption. The problem of contamination is not solved by cleaning the water. The residual nitrate from the cleaning step has to be treated by using bioreactors with denitrifying activity. A viable alternative to treat nitrate contaminated water is to use biological denitrification already existing in aquifers. Denitrification can potentially serve as a biological controlling mechanism for nitrate removal in contaminated groundwater aquifers. The basis for effective self-remediation of aquifers is optimal denitrifying activity under prevailing conditions. Organic carbon has been identified to be the major limiting factor for denitrification in groundwater (Smith and Duff 1988, Bradley et al. 1992, Yeomans et al. 1992, Desimone and Howes 1996) suggesting that reliable prediction of the remediative capability of aquifers would require information on available carbon concentrations of the aqueous and solid phase. The advantages of in situ denitrification, despite the problem with organic carbon limitation, are numerous. The techniques are simple, cost-effective and environmentally sound in comparison to above-groundwater treatment. Other potential beneficial effects of in situ techniques for nitrate removal include simultaneous biodegradation of organic compounds, in some cases toxic pollutants originating from planned or accidental release to aquifers. There is no system without disadvantages and the in situ techniques are no exceptions. Denitrifying activity in groundwater aquifers is suboptimal compared to the optimal conditions created for the techniques used above-ground. The most severe problem is the short supply of organic carbon that serve as electron donor in anaerobic respiration. Several studies using methanol, sucrose, ethanol and solid material (straw) as external carbon sources have reported an increase in denitrification activity ranging from IO to 100% efficiency (Boussaid et al. 1988, Janadaet al. 1988, Mercado et al. 1988, Hamon and Fustec 1991, Dahab and Lee 1992, Weier et al. 1994). The drawback of adding carbon is clogging effects of pore spaces of aquifer sediments. Increase of biomass of microorganisms in combination with gas production (N20 and N2) may, in the worst of cases, block groundwater flow which limits bioremediation efficiency. Another disadvantage is the risk of contamination of the groundwater with pathogenic bacteria. If the water is intended to be used as drinking water this must be addressed as a

21 problem. Stimulating heterotrophic denitrification may be the solution to increased nitrate removal in some areas, but if severe negative effects are the result, then alternative approaches must be explored.

Autotrophic denitrijcation

Autotrophic denitrification may be the answer in areas where organic carbon addition is impractical. Autotrophs use hydrogen gas, sulfur compounds and ferrous iron as electron donors in aerobic and anaerobic respiration. The advantage of such a bioremediative effort is that it is only directed to certain species of bacteria which would lower the increase in biomass and minimise the risk of clogging. Reactor based experiments have shown that autotrophic denitrification is capable of removing considerable amounts of nitrate (Kurt et al. 1987, Se- lenka and Dressler 1990). Basic knowledge of autotrophic denitrification in groundwater environments is limited. Smith et al. (1994) examined the autotrophic hydrogen consumption by denitrifiers in a nitrate contaminated sand and gravel groundwater aquifer at Cape Cod, USA. Sediment cores were collected from an active zone of denitrification. In situ nitrate removal rates in sediment slurries were 4.2 nrnol cm.3 h-' and after 9 days, 34% of the nitrate was con- sumed. Addition of hydrogen and formate significantly increased the nitrate removal rates, 23.1 and 57.8 nmol h', respectively, depleting nearly all of the nitrate in 9 days. Nine strains of hydrogen-oxidising denitrifying bacteria were isolated and eight grew autotrophi- cally on hydrogen with either oxygen or nitrate as electron acceptors. In addition, all of the strains had the ability of heterotrophic growth. The authors concluded that since autotrophic bacteria appear to be common in aquifers and have a potential for high nitrate removal rates, autotrophic denitrification may be as important and widespread as the heterotropic counter- part. The oligotrophic condition in most aquifers potentially favour autotrophic denitrification and thus in situ bioremediative efforts should be directed in this way. For a more detailed discussion about advantages and disadvantages with in situ techniques for nitrate removal the reviews of Hiscock et al. (1991) and Mate$ et al. (1992) are recommended.

Bioreinediation of polluted aquifers with nitrate as electron acceptor

Denitrification has received attention in the field of bioremediation during the last IO years. Most of the laboratory and field work on bioremediation of polluted environments has been performed with aerobic bacteria capable of degrading various organic pollutants. A problem with aerobic biodegradation is the poor solubility of oxygen in water which limits the activity of aerobic degraders. This have been solved to some extent by using hydrogen perox- ide which is soluble in water and slowly breaks down to give free oxygen, but hydrogen per- oxide is toxic to aquifer microorganisms and react with inorganic compounds (Lee et al. 1988, Bewley 1996). Oxygen may quickly be depleted by the activity of aerobic bacteria in polluted aquifers followed by a decrease in degradative activity. Rather than maintaining oxic conditions in aquifers, one may consider the possibility to use alternative electron acceptors to stimulate degradation. Nitrate is highly soluble in water, easily distributed in a polluted aquifer and cheap. Organic compounds that can be degraded with nitrate as the electron acceptor are aromatic hydrocarbons, such as polycyclic aromatic hydrocarbons and nitroaromatic compounds, and chlorinated aliphatic compounds (Berry et at. 1987, Evans and Fuchs 1988, Casella and Payne 1996). Denitrifies have been shown to degrade aromatic hydrocarbons in jet fuel and gasoline contaminated aquifers (Gersberg et al. 1991, Hutchins et al. 1991, Mor- gan et al. 1993, Gersberg et al. 1995).

22 Table 3. Abiotic and biotic sources which may influence the survival of introduced microorganisms into the subsurface. Modified from van Veen et al. 1997 and Berry and Hagedom 1991 Source Factor Effect Abiotic Sediment texture Protection from predation in fine-textured clays Substrate availability, i.e. Organic carbon limitation. Decrease in survivaVactivity electron donors and accep- due to oligotrophic conditions. Electron acceptors com- tors mon in the subsurface Temperature Physiological activity and predation efficiency affected PH Decrease in survivallactivity in acid environments. Op- timal pH range 5-8 for most microorganisms. Selection for acid resistant microorganisms. Inorganic nutrientdtmce Physiological activity affected. Generally, not limiting in elements the subsurface. Pristine aquifers may be low in N and P Toxic pollutants Decrease in survivallactivity. Selection of biodegnda- tive, resistant or tolerant microorganisms Biotic Predation Protozoa graze on bacteria. Population density decreases Competition Competition with adapted indigenous microorganisms. Decrease in survivaVactivity

Biorernediation and microbial release

An alternative in situ technique which have gained attention is the potential to remediate contaminated environments by inoculation of specific strains with degradative capabilities. Introduction of natural strains and genetically engineered microorganisms (GEMS) in agricultural soils has been made to control the activity of plant pathogens or improve soil structure and quality (Wilson and Lindow 1993, van Veen et al. 1997). The major problem of microbial release is the rapid decrease in density followed by low growth of the introduced bacteria. The factors that determines the survival and activity of introduced bacteria in any environment are a complex mixture of biotic and abiotic factors. Two major biotic factors that may influence survival of introduced microorganisms in groundwater are predation by protozoa and competition with indigenous microorganisms adapted to the local conditions. Abiotic factors are numerous and vanes both spatially and temporally. Table 3 shows a summary of biotic and abiotic factors which may affect survival and activity of microorganisms introduced to groundwater aquifers. A typical bioremediative study in a sewage contaminated aquifer combining laboratory and field work was reported by Krumme et al. (1 994). The objective of the study was to intro- duce Pseudomonas sp. B13, which degraded 3-chlorobenzoate (3-CB), into the aquifer and monitor its survival as well as the survival and degradative activity of B13 and the GEM, Pseudomonas sp. B13 FRI (FR 120), which degraded 3-CB and 4-methyl benzoate (4-MB), in aquifer microcosms. Sediment and groundwater from three depths of the aquifer, above the pollutant plume and pristine (DI), in the contaminant plume with reduced oxygen concentrations (D2). and deep within the plume with oxygen concentrations below the detection limit (D3)were used. In DI, a rapid decline in cell numbers of B13 was observed after 4 weeks of injection which was attributed to high densities of grazing protozoa and identified as a bio-

23 logical factor that ultimately may determine the fate of natural and genetically engineered microorganisms introduced to aquifers. Cell numbers in D2 and D3 did not decline to any significant degree during ten weeks and could be detected in low numbers for up to 447 days. The survival of B 13 and FR120 in microcosms was similar to that observed in the field experiments suggesting that survival of FR120 was not altered by the genetic engineering performed on this strain. No degradation of 3-CB and 4-MB by FR120 was observed in microcosms with aquifer sediment from DI and D3, the latter for unknown reasons. Degradation in the sediment from D2 corresponded to the increase in cell numbers by both strains. The efficiency of microbial release may be optimised by selecting bacterial strains with physiological properties that allow colonisation and survival in groundwater aquifers. Inocu- lant density may be maximised to ensure initial survival of the strain. Van Veen et al. (I 997) identified ecological selectivity as a strategy to enhance survival and activity of introduced microorganisms. In ecological selectivity, the selection of inoculant strains is determined by some unique feature of the environment in relation to the inoculant strains. Two methods which have been used are suppression of indigenous soil microorganisms with antibiotics (Bashan 1986, Li and Alexander 1990) and specific substrates suitable for the inoculant strain but unavailable to the majority of the indigenous microorganisms (Nishiyama et at. 1993). The use of antibiotic resistance genes to promote survival of introduced strains is questionable since the spread of antibiotic resistance in natural microbial populations may raise some concern about human health hazards. Growth substrates are, generally, not hazardous as opposed to antibiotic resistance genes. Specificity may be obtained by screening inoculant strains to determine the range of possible substrates. Bioremediation of nitrate contaminated environments with denitrifying strains as the sole effort may be applicable when other treatments are impractical or as complement to the addition of external carbon. The main objective is to increase the density of efficient denitrifiers already present in the bacterial community. The main advantage of using strains isolated from the contaminated aquifer itself is that the strains already are adapted to local environmental conditions. Ecological selectivity may be used to ensure growth of the inoculated strain in the first hand but also by stimulating indigenous denitrifiers. Information from paper I and Il may be useful as an element in a strategy for bioremediation of nitrate contaminated aquifers. De- nitrifiers from the contaminated sites maintained high denitrifying activity down to a nitrate concentration of about 3 mg NO<-N I which may be considered to be a goal concentration of remediation. Even if nitrate concentrations are fluctuating or temporarily lowered, the denitrifying bacteria in aquifers with a long history of nitrate loading may uphold their denitrifying activity, which is crucial for a successful bioremediative effort. Denitrifying activity were carbon limited and carbon amendment increased the denitrification activity about a 2-fold. Optimal nitrate removal in these aquifers would, most likely, require a supply of external carbon sources. The use of denitrifying strains in combination with a specific growth enhancing substrate may prove to be a promising in situ application for bioremediation of the nitrate contaminated aquifers.

Fate of nitrate in contaminated aquifers and fertilised forest soils

Groundwater versusforest soils

Subsurface unconsolidated aquifers are water saturated environments with a solid phase essentially composed of gravel, sand and clay whereas forest soils are a more complex environments composed of different soil profiles of specific appearance and composition (Brady

24 1990, Ghiorse and Wilson 1988). The biological activity in aquifers is generally restricted to bacteria and predators such as protozoa while activity in forest soils is elaborate with complex interactions between roots, higher animals, fungi, protozoa and bacteria (Brady 1990, Ghiorse and Wilson 1988). Aquifer sediments are normally oligotrophic environments with respect to organic carbon and other nutrients. The organic carbon content of the sediment is generally lower than 0. I% with only a small fraction being water soluble and the dissolved organic carbon concentrations less than 10 mg carbon per litre of pore water (Ghiorse and Wilson 1988). Available carbon in forest soils is continuously supplied by decomposition and mineralisation of litter and humus in the upper layers. Deeper down in soil profiles the carbon content decreases as does the microbial activity and biomass. Bacterial numbers in subsurface sediments and groundwater normally range between IO5 and IO7 per gram of sediment whereas numbers in coniferous forest soils range between 10' and 10" per gram of soil. (B%ith et al. 1981, Ghi- orse and Wilson 1985, Bone and Balkwill 1988, Pedersen and Ekendahl 1990, Hazen et al. 199 I, paper IV). Nitrate leaching from agricultural soils is considered to be a serious environmental problem today and numerous studies have been performed to elucidate the fate of nitrate in soils, in the vadose zone and in the subsurface (Bergstrom and Johansson 1991, Exner et al. 1991, Geyer et al. 1992, Lowrance 1992, Jones and Schwab 1993, Izaurralde et al. 1995). Nitrate leaching from agricultural soils has been identified as the main source of nitrate contamination of groundwater (Weil et al. 1990, Jemison and Fox 1994). The increase of the nitrate in the subsurface environment is likely due to the mobility of nitrate in agricultural soils. Artificial irrigation and natural precipitation leads to a nitrate transport through the vadose zone to the underlying groundwater. If the nitrate is not assimilated by vegetation, immobilised in microorganisms or denitrified in surface soils while transported, nitrate will accumulate in groundwater and may be transported to sensitive areas such as drinking water wells. Denitrification is the main biological process that can regulate subsurface discharge of nitrate from groundwaters to surface waters. An intriguing question is if denitrification in nitrate contaminated groundwater aquifers may limit the nitrate influx to lakes and coastal waters. In paper I, average denitrification rates for strains from the nitrate contaminated sites were twice as high as rates of the strains from the pristine site. Denitrification were carbon limited and glucose amendment increased the denitrification activity about a 2-fold for all strains, The strain specific differences in denitrification rates increased to a 2.5-fold after carbon addition indicating that the differences in reduction rates cannot be explained by different carbon utilisation rates but rather reflect innate differences in the reductases of the strains. In paper E, the significant differences in denitrifying activity were maintained in a nitrate concentration gradient ranging from 0.5 to 48 NOj--N 1.'. The increased nitrate removal rates in nitrate contaminated aquifers have the potential to limit the nitrate transport, at least as long available carbon concentration does not limit the activity of denitrifies. Most coniferous forests are considered to be nitrogen limited while microbial activity in general are assumed to be carbon limited (Tamm 1991, Foster et al. 1980, Heinrich and Haselwandter 1991). Most soil nitrogen is bound to organic matter or immobilised in micro- . n organisms and plants. The amount of nitrogen that is available for plants and microorganisms , is seldom larger than 1% (Gosz 1981). The main source of inorganic nitrogen in coniferous forest soils is ammonium since the acidic conditions ofthe soils with pH varying between 3.5 and 4.5, which tends to inhibit nitrification of the soils (Johnsrud 1978, Persson and Wir6n 1995). Fertiliser experiments in forest soils have shown considerable positive effects on plant growth (Tamm 1991, Martikainen 1996). Nitrogen limited forest ecosystems appear to accumulate increased nitrogen loads from atmospheric deposition or fertilisation simply by a plant growth response. However, increasing atmospheric deposition of nitrogen has been shown to have adverse effects on forest vegetation, promote soil acidification, heterotrophic microbial activity and leaching of nitrate and base cations (Aber et al. 1989, Schulze et al. 1989, Stams et al. 1991, Martikainen 1996). If the nitrogen load from deposition, fertilisation, and nitrification reach the level where organic matter, vegetation and microorganisms can no longer bind or assimilate the nitrogen, the soil may become nitrogen saturated and start to leach nitrogen compounds. An intriguing question is if denitrification in fertilised forest soils may counteract the increase of nitrate. The role of denitrification as a nitrogen loss mechanism in undisturbed and in clear-cut forest soils is considered to be limited and in fertilised forest soil, large variation in denitrification rates appear to be independent of the nitrate concentration (Gundersen 1991, Martikainen 1996). In paper IV, denitrifying activity was negatively correlated with the amount of fertiliser applied to the soil, Le. the denitrification flux was highest in the control soil and lowest in the soil that had received the highest amount of fertiliser. This indicates that denitrification in the fertilised forest soils cannot compensate for the excess of nitrate. DNRA dominated over denitrification in all soils while nitrification was low and there were no obvi- ous relationship between ammonium produced and the amount of fertiliser applied. The dominant fate of added nitrogen in our study was immobilisation (between 64 and 97%), but fertilised soils immobilised less of the added inorganic nitrogen than the control soil, which indicate that the fertilised soils are becoming nitrogen saturated Similarly, nitrogen fertilisation reduced the immobilisation of nitrogen in microorganisms in fertilised forest soils in Finland (Priha and Smolander 1995). Fertilisation does not appear to be a basis for selection of nitrogen transformers in coniferous forest soils. Abiotic and biotic immobilisation and mo- bilisation of nitrogen seem to be quantitatively more important processes in these soils.

C/N ratio determining the fare of nitrate

How do the differences in C/N ratios affect nitrogen transformation in a nitrate contaminated groundwater environment versus a fertilised coniferous forest soil environment? The two major nitrate reducing processes, denitrification and DNRA, are similar processes, both use nitrate and both occur in anoxic environments. Denitrification is an energy generating process as well as the first enzymatic step in DNRA. The question is in which environment does one or the other dominate and under what environmental conditions. Tiedje et al. (1982) suggested that the major factor controlling the competition between denitrifiers and ammonifiers is the C/N ratio. In environments such as groundwaters, which are oligotrophic with respect to the available carbon but not to the nitrate of the interstitial water (low CM ratio), denitrification is expected to the dominate over DNRA. Studies performed with estaurine sediments and groundwater sediments low in carbon but with high nitrate concentrations confirm this suggestion (King and Nedwell 1987, Smith and Duff 1987, Binnerup et al. 1992). In paper I, a small proportion (9-22%) of isolated strains reduced nitrate to nitrite. This group may, partially, be represented by strains with DNRA activity. The low frequency of potential DNRA strains supports the idea that environments with low C/N ratios promotes denitrification rather than DNRA. In environments rich in carbon and low in nitrate such as the rumen, digested sludge and anaerobic sediments (high C/N ratio), DNRA is quantitatively the most important nitrate reducing process (Sprrensen 1978, Kaspar et al. 1981, Tiedje et al. 1982, Tiedje 1988). DNRA may be the favoured process in coniferous forest soils since the CM ratio, generally, range from 20 to 40 (Priha and Smolander 1995, Persson and Wirh 1995, Koopmans et al. 1995, paper UI). In our study the DNRA activity was quantitatively the most important biological nitrogen transforming process in the forest soils (paper IV). However,

26 deviating results have been reported by Bengtsson and Annadotter (1989) who investigated denitrification and DNRA activity of a groundwater aquifer in southern Sweden using oxic and anoxic groundwater continuos-flow columns. By adding "NO? and tracing the fate of the isotope, they determined that 80-90% of the "NOi in the oxic column was reduced to nitrogen gas and IO-20% to ammonium. In the anoxic column, 35% of the nitrate was reduced to nitrogen gas and 50% to ammonium. Judging from the results, DNRA appear to occur in environments with low CN ratios and may under anoxic conditions even dominate over denitrification. As a consequence, it may pose a serious problem if carbon concentrations increase in nitrate contaminated aquifers, which may select for DNRA strains rather than for denitrifying strains. Nitrate will then to a higher degree be conserved as ammonium in the aquifers. Nitri- fying bacteria may oxidise the ammonium to nitrate, which partly may be transported with the groundwater and partly be reused by DNRA strains. The problem may further be accentuated by the fact that typical in situ bioremediative applications rely on external carbon to boost denitrification.

Future perspectives and research

Groundwater is important for all of us because it is the basis for much of our drinking water supply and the concern about groundwater quality world wide is growing steadily. The cause is known to us. Accidental or planned release of organic or inorganic pollutants, eventu- ally, reach the groundwater with variable degrees of contamination as a result. In the case of nitrate contamination, specific denitrifying strains for bioremediative purposes are an exciting concept that merit further study. The denitrifiers have so far been characterised in single strain experiments to elucidate whether phenotypic characters such as denitrifying activity reflect potential genotypic differences between denitrifying strains isolated from aquifers varying in nitrate concentration. It appears that denitrification is an adaptive trait in some nitrate contaminated environments where nitrate is easily accessible to the denitrifiers which is the case for groundwater aquifers. The information on denitrifying activity is only a first step. For a proper application, further characterisation of the denitrifying strains will be required and should include basic knowledge of survival and competitive abilities for substrates other than nitrate in the presence of indigenous microorganisms in laboratory and field experiments. The concept of ecological selectivity is a key factor that should be explored in detail. The ultimate aim is to use the basic knowledge to be able to initiate full-scale bioremediation of nitrate contaminated groundwater aquifers Molecular biology offer a number of sophisticated techniques to determine gene and protein structure which may be used to identify and characterise the molecular target of adaptation in strains from the nitrate contaminated aquifers. Reductase DNA sequences from different bacteria can be acquired to produce probes for detection of reductase genes in the groundwater strains. Gene cloning accompanied by DNA sequencing may be a useful tool to determine whether the denitrifying strains from the nitrate contaminated aquifers contain duplicated genes as identical copies or genes that codes for enzymes with different activity or affinity. The presence of plasmids containing reductase genes may be explored to elucidate whether horizontal processes are a potential mechanism for genetic adaptation in denitrifiers.

Acknoivledgentenls I am grateful to Goran Bengtsson for helpful comments on previous drafts. The work included in the thesis was supported by grants from the Swedish Environmental Protection Board. Swedish Natural Science Research Council and Crafoordska Stiftelsen

21 References

Aber J.D.. Nadelhoffer K.J., Steudler P.A. Melillo J.M. (1989) Nitrogen saturation in forest ecosystems. Bioscience 39:378-386. Anonymous 1980. European Economic Community (EEC). Council directive on the quality of water for human consumption. - Official J. No 80I778EECL. 229:11-29. Arai H.. Igarashi, Y.. Kcdama T. (1991) Nitrite activates the transcription of the fseumottas aeruginosa nitrite reductase and cytochrome c-551 operon under anaerobic conditions. FEBS Lett. 288:227-228. Avcrill B. A. (1996) Dissimilatory nitrile and nitric oxide reductases. Chem. Rev. 96:2951-2964. Bashan Y. (1986) Enhancement of wheat root colonization and plant development by Azopirillitm brasiliense Cd. following temporary depression of rhizosphere microflora. Appl. Environ. Microbiol. 51:1067-1071. Beauchamp E.G.. Trevors J.T., Paul J.W. (1989) Carbon sources for bacterial denitrification. In: Adv. Soil Sci. Vol IO. Springer Verlag, New York., pp. 113-142. Bell L.C., Richardson D.J.. Ferguson S.J. (1990) Periplasmic and membrane-bound respiratory nitrate reductases in Thiosphaera panrorropha. FEBS Lett. 265%-87. Bell L.C., Ferguson J. (1991) Nitric and nitrous oxide reductases are active under aerobic conditions in cells of Thiosphaera panrotropha. Biochem. J. 273:423-427. Bengtsson G, Annadotter H. (1989) Nitrate reduction in a groundwater microcosm determined by "N gaschromatography-masspectrometry.AppLEnviron. Microbiol. 55:2861-2870. Bergstrom L., Johansson R. (1991) Leaching of nitrate from monolith lysimeters of different types of agricultural soils. J. Environ. Qual. 202301-807. Berks B.C.. Baratta D., Richardson D.J., Ferguson S.J. (1993) Purification and characterization of a nitrous oxide reductase from Thiosphaera panrorropha. Implications for the mechanism of aerobic nitrous oxide reduction. Eur. J. Biochem. 212:467-476. Berks B.C.. Richardson D.J.. Robinson C., Reilly A., Aplin R.T.. Ferguson S.J., (1994) Purification and characterization of the periplasmic nitme reductase from Thiosphaera panrorropha. Eur. J. Biochem. 220I 17- 124. Berks B.C.. Ferguson S.J.. Moir J.W.B. Moir. Richardson D.J. (1995) Enzymes and associated electron transport systems that can catalyse the respiratory reduction of nitrogen oxides and oxyanions. Biochim.Biophys. Acta 1232:97-173. Berry D.F.. Francis A.J., Bollag J-M. (1987) Microbial metabolism of homocyclic and heterocyclic aromatic compounds under anaerobic conditions. Microbiol. Rev. 5 I :43-59. Berry D.F.. Hagedorn C. (1991) Soil and groundwater transport of microorganisms. In: Assessing ecological risks of biotechnology. (Ginzburg L.R. Ed.). Butterworth-Heineman. Stoneham. pp. 57-73 Bewley R.J.F. (1996) Field implementation of in situ bioremediation: key physicochemical and biological factors. In: Soil biochemistry vol. 9 (Stotzky G.. Bollag J-M. Eds.). Marcel Dekker Inc.. New York, pp. 473-541. Binnerup S.J., Jensen K.. Rcvsbech N.P., Hjonh Jensen M., SBrensen J. (1992) Denitrification, dissimilatory reduction of nitrate lo ammonium, and nitrification in a bioturbated estuarine sediment as measured with "Nand microsensor techniques. Appl. Environ. Microbiol. 58:303-313. Blasco F., Nunzi F., Pommier J., Brasseur R., Chipaux M.. Giordano G. (1992) Formation of active heterologous nitrate reductases between nitrate reductase A and 2 of Escherichia coli. Mol. Microbiol. 6~209-219. Bock E., Schmidt I.. Stueven R., Zart D. (1995) Nitrogen loss caused by denitrifying Nitrosoinonas cells using ammonium or hydrogen as electron donors and nitrite as electron acceptor. Arch. Microbiol. 163: 16-20, Bone T.L.. Balkwill D.L. (1988) Morphological and cultural comparison of microorganisms in surface soils and subsurface sediments at a pristine study site in Oklahoma. Microb. Ecol. 16:49-64. Bonin P.. Gilewicz M. (1991) A direct demonstration of 'co-respiration' of oxygen and nitrogen oxides by fseiidonionas naurica: some spectral and kinetic properties of the respiratory components. FEMS Microbiol. Lett. 80183- 188. Bouissad F., Martin G.,Morvan J.. Collin J.J.. Landreau A., Talbo H. (1988) Denitrification in-situ of groundwaters with solid carbon matter. Environ. Technol. Lett. 95303-8 16. Bradley P.M., Fernandez J.R., Chapelle F.H. (1992) Carbon limitation of denitrification in an anaerobic groundwater system. Environ. Sci. Technol. 262377-2381. Brady N.C. (1990)The nature and property ofsoils. IO ed. Macmillan Publishing Company. New York.

28 Bruning-Fann C.S.. Kaneene J.B. (l993a) The effects of nitrate. nitritc and n-nilroso compounds on animal health: a review. Vet. Hum. Toxicol. 35:237-253. Bruning-Rnn C.S.. Kancene J.B. (l993b) The effects of nitrate. nitrite and n-nitroso compounds on human health: a review. Vet. Hum. Toxicol. 35:521-538. Baath E., Lundgren B.. Soderstrom B. (1981) Effects of nitrogen fertilization on the activity and biomass of fungi and bacteria in a podzolic soil. Zentralblatt Bakteriol.. Mikrobiol. Hygiene I. Abt. Orig. C290-98. Calder K.. Burke K.A.. Lascelles J. (1980) Induction of nitrate reductase and membrane cytochromes in wild type and chlorate resistant Paracoccus denirrificarts.Arch. Microbiol. 126 149- 153. Capone D.G.. Bautista M.F. (1985) A groundwater source of nitrate in nearshore marine sediments. Nature 313:214-216. Carter J.P., Hsin Hsiao Y.. Spim S.. Richardson D.J. (1995) Soil and sediment bacteria capable of aerobic nitrate reduction. Appl. Environ. Microbiol. 61,2852-2858. Casella S.. Paync W.J. (1996) Potential of denitrifiers for soil environment protection. FEMS Microbiol. Lett. 1401-8. Castignetti D.. Hollocher T.C. (1984) Heterotrophic nitrification among denitrifiers. Appl. Environ. Microbiol. 47:620-623. Chan Y-K.. Wheatcroft R. (1993) Detection of a nitrous oxide reductase structural gene in Rhizobium nielilori strains and its location on the nod megaplasmid of JJlclO and SU47. J. Bacteriol. 175: 19-26. Clarke P.H. (1983) Evolution of new phenotypes. In: Current perspectives in microbial ecology. (Klug M.J., Rcddy C.A. Eds.) Proceedings of the third international symposium on microbial ecology. Michigan State University. 7-12 August. ASM. Washington DC, pp. 71-78. Cole J.A. (1990) Physiology, biochemistry and genetics of nitrate dissimilation lo ammonia. In: Denitrification in soil and sediment. (Rcvsbech N.P.. Sorensen J.. Eds.) Fems Symposium No. 56. Plenum Press, Ncw York, pp. 57-76. Cole J.A. (1996) Nitrate reduction to ammonia by enteric bacteria: redundancy, or a strategy for survival during oxygen starvation? FEMS Microbiol. Ecol. 1361-1 1. Coyne M.S.. Arunakumari A,. Averill B.A.. Tiedje J.M. (1989) Immunological identification and distribution of dissimilatory heme cd, and nonheme copper nitrite reductase in denitrifying bacteria. Appl. Environ. Microbiol. 55:2924-2931. Coyne M.S.. Tiedje J.M. (1990a) Distribution and diversity of dissimilatory NO; reductases in denitrifying bacteria. In: Denitrification in soil and sediment. (Revsbech N.P.. Sorensen I., Eds.) Fems Symposium No. 56. Plenum Press, New York, pp. 21-35. Coyne M.S.. Tiedje J.M. (1990b) Induction of denitrifying enzymes in oxygen-limited Achromobacrer cycloclasres continuous culture. EMSMicrobiol. Ecol. 73:263-270. -. Dahab M.F.. Lee P.Y. (1992) Nitrate reduction by in-situ biodenitrification in groundwater. Wat. Sci. Technot. 26:1493-1502. Denariaz G.. Payne W.J., Le-Gall J. (1989) A halophilic denitrifier Bacillus hnlodertirrificans new species. 1- Int. J. Sysl. Bacteriol. 39:145-151. Desimone L.A.. Howes B.L. (1996) Denitrification and nitrogen transport in a coastal aquifer receiving wastewater discharge. Environ. Sci. Technol. 30: I 152- 1162. Evans C.H.. Fuchs G.(1988) Anaerobic degradation of aromatic compounds. Ann. Rev. Microbiol. 42:289-317. Exner M.E.. Burbach M.E., Watts D.G.. Shcarman R.C.. Spalding R.F. (1991) Deep nitrate movement in the unsaturated zone of a simulated urban lawn. J. Environ. Qual. 20658-662. Ferguson S. (1994) Denitrification and its control. Antonie van Leeuwenhoek 66:89- I IO. Foster N.W.. Beauchamp E.G.. Corke C.T. (1980) Microbial activity in a Pinus bartkriann forest floor amended with nitrogen and carbon. Can. J. Soil Sci. 60.199-210 Freitag A.. Ruden M., Bock E. (1987) Growth of Nirrobacfer by dissimilatory nitrate reduction. FEMS Microbiol. Lett. 48:105-109. Pried J.J. 1991. Nitrates and their control in the EEC aquatic environment. - In: Bogardi. I. and Kuzelka R. D. (eds.) Nitrate contamination: Exposure, consequence. and control. NATO AS1 Ser. G: Ecological Sciences 30. Springer-Verlag. Berlin, pp. 3-1 1. Fuggi. A.. Vona V., Di Martino Rigano V., Di Martino. C.. Martello A.. Rigano C. (1984) Evidence for two transport systems for nitrate in the acidophilic, thermophilic alga Cyariidium caldarium. Arch. Microbiol. 137:281-285. Gersbcrg R.M., Dawsey W.J.. Bradley M.D. (1991) Biodegradation of monoaromatic hydrocarbons in groundwater under denitrifying conditions. Bull. Environ. Contam. Toxicol. 47:230-237. Gersbcrg R.M.. Konh K.G.. RiccL.E.. Randall J.D. Bogardt A.H.. Dawsey W.J.. Hemmingsen B.B. ( 1995) Chemical and microbial evaluation of in-sin1 biorcmediation of hydrocarbons in anoxic groundwater enriched with nutrients and nitrate. W. J. Microbiol. Biotechnol. I1:549-558.

29 Geyer D.J.. Keller C.K.. Smith J.L.. Johnstone D.L. (1992) Subsurface fate of nitrate as a function of depth and landscape position in Missouri Flat Creek watershed, U.S.A. J. Contam. Hydrol. I 1:127- 147. Ghiorse W.C., Wilson J.T. (1985) Characterization of subsurface bacteria associated with two shallow aquifers in Oklahoma. Appl. Environ. Microbiol. 50580-588. Ghiorse W.C., Wilson J.T. (1988) Microbial ecology of the terrestrial subsurface. Adv. Appl. Microbiol. 33:107-172. Gosz J.R. (1981) Nitrogen cycling in coniferous forests. In: Terrestrial nitrogen cycles. (Clark F.E., Rosswall T. Eds.). Ecol. Bull. (Stockholm) 33:405-426. Gundersen P. (1991) Nitrogen deposition and the forest nitrogen cycle: role of denitrification. For. Ecol. Manag. 44: 15-28 Hamon M., Fustec E. (1991) Laboratory and field study of an in-situ groundwater denitrification reactor. Res. J. Water. Pollut. Control Fed. 63942-949. Hazen T.C.. JimenCz, L. L6pez de Victoria G., Fliermans C.B. (1991) Comparison of bacteria from deep subsurface sediment and adjacent groundwater. Microb. Ecol. 22:293-304 Henrich M.. Haselwandter K. (1991). Denitrification and enzyme activity in a norway spruce forest. For. Ecol. Manage. 44:63-68. Hernandez D., Rowe J.J. (1988) Oxygen inhibition of nitrate uptake is a general regulatory mechanism in nitrate respiration. J. Biol. Chem. 263:7937-7939. Hirsch P., Rades-Rohkohl E. (1988) Some problems in the determination of viable counts of groundwater microrganisms. Microb. Ecol. 16:99-113. Hiscock K.M. Lloyd J.W.. Lerner D.N. (1991) Review of natural and artificial denitrification of groundwater. Wat. Res. 25: 1099- I 1 I I. Ho T.P.. Jones A.M., Hollocher T.C. (1993) Denitrification enzymes of Bacillus srenrorherntophil~~s.FEMS Microbiol. Lett. I14:135-138. Hochsrein L.I., Tomlinson G.A. (1988) Thc enzymes associated with denitrification. Ann. Rev. Microbiol. 4223 1-261. Hollocher T.C.. Kristjhsson J.K. (1992) Termophilic denitrifying bacteria: A survey of hot springs in south-western Iceland. FEMS Microbiol. Ecol. 101:113-1 19. Hooper A.B., Vannelli T.,Bergmann D.J. Arciero D.M. (1997) Enzymology of the oxidation of ammonia to nitrile by bacteria. Antonie van Leeuwenhoek 7159-67. Hutchins S.R., Sewell G.W., Kovacs D.A.. Smith G.A. (1991) Biodegradation of aromatic hydrocarbons by aquifer microorganisms under denitrifying conditions. Environ. Sei. Technol. 25:68-76. lwurralde R.C.. Feng Y.. Robertson J.A.. McGill W.B.. JumaN.G. Olson B.M. (1995) Long-term influence of cropping systems, tillage methods and N source on nitrate leaching. Can. J. Soil Sci. 75:497-505. Janada V.. Rudovsky J., Wanner J., Marha K. (1988) In situ denitrification of drinking water. War. Sci. Technol. 20:215-220. Jemison J.M.. Fox R.H. (1994) Nitrate leaching from nitrogen fenilised and manured corn measured with zero-tension pan lysimeter. J. Environ. Qual. 23:337-342. Jetten M.S.M., Longemann S. Muyzer G. Robertson L.A.. de Vries S., van Loosdrecht M.C.M., Kuenen J.G. (1997) Novel principles in the microbial conversion of nitrogen compounds. Antonie van Leeuwenhoek 7 I :75-93. Johannes R.E. (1980) The ecological significance of the submarine discharge of groundwater. Mar. Ecol. Prog. Ser. 3:365-373. Johnsrud S.C. (1978). Heterotrophic nitrification in acid forest soils. Holarc. Ecol. 1:27-30. Jones R.D., Schwab B.P. (1993) Nitrate leaching and nitrite occurrence in a fine-textured soil. Soil Sci. 155:272-282. Karsten G.K.. DrakeH.L. (1997) Denitrifying bacteria in the earthworm gastrointestinal tract and in vivo emission of nitrous oxide (N20) by earthworms. Appl. Environ. Microbiol. 63: 1878-1882. Kaspar H.F.. Tiedje J.M. Firestone R.B. (I98 I) Denitrification and dissimilatory nitrate reduction to ammonium in digested sludge. Can. J. Microbiol. 27378-885. Kennedy M.J.. Lawless J.G. (1985) Role of chemotaxis in the ecology of denitrifiers. Appl. Environ. Microbiol. 49:109-1 14. King D.. Nedwell D.B. (1987) The adaptation of nitrate-reducing communities in estaurine sediments in response to overlying nitrate load. EMSMicrobiol. Ecol. 4215-20. Klemedtsson L.. Hansson GB.. Mosier A. (1990) The use of acetylene for the quantification of N and N20 production from biological processes in soil. In: Denitrification in soil and sediment. (Revsbeeh N.P., Sorensen J. Fems Symposium No. 56. Plenum Press. New York. pp. 167-180. Knowles R. (1982)as.) Denitrification. Microbiol. Rev. 4643-70. Knowles R. (1990) Acetylene inhibition technique: development. advantages. and potential problems. In:

30 Denitrification in soil and sediment. (Revsbech N.P.. S~rensenJ. Us.) Fems Symposium No. 56. Plenum Press. New York. pp. 151-166. Koopmans C.J.. Lubrecht W.C.. Tietema A. (1995) Nitrogen transformation in two nitrogen saturated forest ecosystems subjected to an experimental decrease in nitrogen deposition. Plant and Soil 174:205-218. Korom S.F. (1992) Natural denitrification in the saturated zone: a review. Water Resour. Res. 28: 1657-1668. Krumme M.L., Smith R.L.. Egestorff J., Thiem M.. Tiedje J.M., Timmis K.N., Dwyer D.F. (1994) Behavior of pollutant-degrading microorganisms in aquifers: predictions for genetically engineered organisms. Environ. Sci. Technol. 281 134-1 138. Kuenen.J.G., Robertson L.A. (1994) Combined nitrification-denitrification processes. FEMS Microbiol. Rev. 15:109-117. Kurt M., Dunn. I.J.. Bourne. J.R. (1987) Biological denilrification of drinking water using autotrophic organisms with hydrogen in a fluidized-bed biofilm reactor. Biotechnol. Bioeng. 29:493-501. Karner H., Zumft W.G. (1989) Expression of denitrification enzymes in response to the dissolved oxygen level and respiratory substrate in continuos culture of Pseudomonas srurzeri. Appl. Environ. Microbiol. 55:1670-1676. Lammel G.. Grass1 H. (1995) Greenhouse effect of NO-x. Environ. Sci. Poll. Res. Int. 2:40-45. Li D-M.. Alexander M. (1990) Factors affecting co-inoculation with antibiotic-producing bacteria to enhance rhizobia1 colonization and nodulation. Plant and Soil 129:195-201. Lee M.D., Thomas J.M., Borden R.C.. Bedient P.B., Wilson J.T.. Ward C.H. (1988) Biorestoration of aquifers contaminated with organic compounds. Crit. Rev. Environ. Control I8:29-89. Lloyd D., Boddy L., Davies K.J.P.(1987) Persistence of bacterial denitrification under aerobic conditions: the rule rather than the exception. EMSMicrobiol. Ecol. 45: 185-190. Lowrance R. (1992) Nitrogen outputs from a field-size agricultural watershed. J. Environ. Qual. 21:602-607. Martikainen P.J. (1996) Microbial processes in boreal forest soils as affected by forest management practices and atmospheric stress. In: Soil biochemistry vol. 9 (Stotzky G.. Bollag I-M. Eds.) Marcel Dekker Inc.. New York, pp.195-232. MatEjii V.. Ciiinskfi S., Krejfi J., Janoch T. (1992) Biological water denitrification-a review. Enzyme Microbiol. Technol. 14:170-183. McNaughton SJ.. Wolf L.L. (1979) General Ecology. Holt. Rhinehart and Winston, New York. Mercado A.. Libhaber M.. Soares M.I.M. (1988) In situ biological groundwater denitrification: concepts and preliminary field tests. Wat. Sci. Technol. 20197-209. Moir J.W.B.. Baratta D.. Richardson D.J., Ferguson S.J. (1993) The purification of a cdl-tpe nitrite reductase from, and the absence of a copper-type nitrite reductase from, the aerobic denitrifier Thiosphaera panrorropha; the role of pseudoazurin as an electron donor. Eur. J. Biochem. 212377-385. Morgan P., Lewis S.T.. Watkinson R.J. (1993) Biodegradation of benzene, toluene. ethylbenzene and xylenes in gas-condensate-contaminated groundwater. Environ. Poll. 82:181- 190. Nishiyama M.. Senoo K.. Matsumoto S. (1993) Establishment of y- I .2.3.45,6-hexachlorocyelohexane assimilating bacterium. Sphingomonaspaucimobilisstrain SS86. in soil. Soil Biol. Biochem. 25769-774. Noji. S.. Taniguchi S. (1987) Molecular oxygen controls nitrate transport of Escherichia coli nitrate- respiring cells. J. Biol. Chem. 2629441-9443. Parkin T.B.. Sextone A.J., Tiedje J.M. (1985) Adaptation of denitrifying populations to low soil pH. Appl. Environ. Microbiol. 491053-1056. Pateureau D.. Davison J.. Bernet N., Moletta R. (1994) Denitrification under various aeration conditions in Contanionas sp.. strain SGLY2. FEMS Microbiol. Ecol. 14:71-78. Pedersen K.. Ekendahl S. (1990) Distribution and activity of bacteria in deep granitic groundwaters of southeastern Sweden. Mierob. Ecol. 20:37-52. Persson T., Wiren A. (1995) Nitrogen mineralization and potential nitrification at different depths in acid forest soil. Plant and Soil 168-169:55-65. Pigliucci M. (1996) How organisms respond to environmental changes: from phenotypes to molecules (and vice versa). Trends Ecol. Evol. I1:168-173. Poth M.. Focht D.D. (1985) "N kinetic analysis of N20 production by Nirrosomonas europaea: an examination of nitrifier denitrification. Appl. Environ. Mierobiol. 49: I 134- 1 14 I. Poth M. (1986) Dinitrogen production from nitrite by Nirrosomonns isolate. Appl. Environ. Microbiol. 52957-959, Powers J.F., Schepers J.S. (1989) Nitrate contamination of groundwater in North America. Agric. Ecosys. Environ. 26165-187. Priha 0.. Smolander A. (1995) Nitrification. denitrification and microbial biomass N in soil from two N-fertilized and limed Norway spruce forests. Soil. Biol. Biochem. 27:305-3 IO. Prosser 1.1.. Cox D.J. (1982) Nitrification. In: Experimental microbial ecology. (Bums R.G..

31 Slater H.J. Eds.) Blackwell Scientific Publications, Oxford, pp.178-193. Rees E., Siddiqui R. A,, Koster F., Scheneider B.. Friedrich B. (1997) Structural gene (riirS) for the cytochrome cdl nitrite reductase of Alicaligeries eurrophirs HI 6. Appl. Environ. Microbiol. 632300-802. Remde A.. Conrad R. (1990) Production of nitric oxide in Nirrosomorias europaea by reduction of nitrite. Arch. Microbiol. 154:187-191. Richardson D.J., Ferguson S.J. (1992). The influence of carbon substrate on the activity of the periplasmic nitrate reductase in aerobically grown Thiosphaera paiirorropha. Arch. Microbiol. 157535537. Robenson K. (1991) Emissions ofN2O in Sweden-natural and anthropogenic sourecs. Ambio 20, 151-155. Robertson L.A.. Kuenen J.G. (1983) Thiosphaera parirorropha gen. nov. sp. nov., a facultatively anaerobic, facultatively autotrophic sulphur bacterium. J. Gen. Microbiol. 129: 2847-2855. Robertson L.A.. Kuenen J.G. (1984) Aerobic denitrification: a controversy revived. Arch. Microbiol. 139:351-354. Robertson L.A., Kuenen J.G. (1992) The effect of electron acceptor variations on the behaviour of Thiosphaera paiirorropha and Paracoccus deriirrificarrsin pure and mixed cultures. FEMS Microbiol. Ecol. 86:221-228. Rozak D.B., Colwell R.R. (1987) Survival strategies of bacteria in the natural environment. Microbiol. Rev. 5 1:365-379. Rijmermann D., Friedrich B. (1985) Denitrification by Alicaligeties eurrophits is plasmid dependent. J. Bacterid. 162852-854. Saad O.A.L.A., Conrad R. (1993) Adaptation to temperature of nitric oxide-producing nitrate-reducing bacterial populations in soil. Sysl. Appl. Microbiol. 16:120-125. Schlichting C.D., Pigliucci M. (1993) Control of phenotypic plasticity via regulatory genes. Amer. Nat. 142366-370. Schneider B., Nies A.. Friedrich B. (1988) Transfer and expression of litoautotrophy and denitrification in a host lacking these metabolic activities. Appl. Environ. Microbiol. 54:3 173-3 176. Schulze E-D., De Vries W., Hauhs M., RosCn K., Rasmussen L.. Tamm C.O.. Nilsson I. (1989) Critical loads for nitrogen deposition on forest ecosystems. Water, Air and Soil Poll. 48:451-456. Selenka F., Dressler R. (1990) Microbiological and chemical investigations on a biological autotrophic denitrification plant using hydrogen as an energy source. Aqua 39: 107-1 16. Shieh W.Y. Liu C.M. (1996) Denitrification by a novel halophilic fermentative bacterium. Can. I. Microhiol. 42507-5 14. Shoun H., Kim D.H. Uchiyama H., Sugiyama J. (1992) Denitrification by fungi. FEMS Microbiol. Leu. 94:217-281. Siddiqui R.A., Warneke-Eberz U., Hengsberger A. Schneider B.. Kostka S., Friedrich B. (1993) Structure and function of a periplasmic nitrate reductase in Alicaligeries eittrophrts H16. J. Bactcriol. 1755867-5876. Smith R.L.. Duff J.H. (1988) Denitrification in a sand and gravel aquifer. Appl. Environ. Microbiol. 54: 1071-1078. Smith R.L.. Ceazan M.L.. Brooks M.H. (1994) Autotrophic, hydrogen-oxidizing, denitrifying bacteria in groundwater, potential agents for bioremediation of nitrate contamination. Appl. Environ. Microbiol. 60: 1949-1955. Spalding R.F., Exner M.E. (1993) Occurrence of nitrate in groundwater-a review. I. Environ. Qual. 22392402, Stams A.J.M.. Booltink H.W.G.,Lutke-Schipholt I.J., Beemsterboer B.. Woittiez J.R.W., van Breemen N. (1991) A field study on the fate of "N-ammonium to demonstrate nitrification of atmospheric ammonium in an acid forest soil. Biogeochem. 13:241-25 Staver K.W.. Brinsfield R.B. (1996) Seepagc of groundwater nitrate from a riparian agroecosystem into the Wye River estuary. Estuaries 19359-370. Stouthamer A.H. (1988) Dissimilalory reduction of oxidiscd nitrogen compounds. In: Biology of anaerobic microorganisms. (Zehnder A.J.B. Ed.) Wiley, New York, pp. 245-303. Sclrensen J. (I 978) Capacity for denitrification and reduction of nitrate to ammonia in a coastal marine sediment. Appl. Environ. Microbiol. 35301-305. Tamm C.O. (1991) Nitrogen in the terrestrial environment. Questions of productivity, vegetational changes and ecosystem stability. Ecological studies 8 I. Springer Verlag. Berlin. I 16 pp. Thayer J. R.. Huffaker R.C. (1982) Kinetic evaluation, using I3N, reveals two assimilatory nitrate transport systems in Klebsiella piiemoiiiae. J. Bactcriol. 149: 198-202. Tiedje J.M.. SBrensen J. Chang Y.Y.L. (1981) Assimilatory and dissimilatory nitrate reduction: perspectives and methodology for simultaneous measurement of several nitrogen cycle processes. In: Terrestrial nitrogen cycles. (Clark F.E.. Rosswall T. Eds.). Ecol. Bull. (Stockholm) 33:331-342. Ticdje J.M.. Sexstone A.J.. Myrold D.D.. Robinson J.A. (I 982) Denitrification: ecological, competition and survival. Antonie van Leeuwenhoek J. Microbiol. 48569-583.

32 Tiedje J.M. (1988) Ecology of denitrification and dissimilatory nitrate reduction to ammonium. In: Biology of anaerobic microorganisms. (Zehnder A.J.B. Ed.) Wiley, New York. pp. 179-244. Tomlinson G.A.. Jahnke L.L., Hochstein L.I. (1986) Halobacterium denirrificarrs new species an extremly halophilic denitrifying bacterium. Int. J. Syst. Bacteriol. 3666-70. Van der Meer J.R.. de Vos W.. Harayama S.. Zehnder A.J.B. (1992) Molecular mechanisms of genetic adaptation to xenobiotic compounds. Microbiol. Rev. 56677-694. Van Veen J.A.. van Overbeek L.S.. van Elsa J.D. (1997) Fate and activity of microorganisms introduced intosoil. Microbiol. Mol. Rev. 61:121-135. Warneke-Ebcrz U.. Friedrich B. (1993) Three nitrate reductase activities in Alcaligenes ercrrophns. Arch. Microbiol. 159:405-409. Weier K.L., Doran J.W.. Mosier A.R., Powe J.F., Peterson T.A. (1994) Potential for bioremediation of high nitrate irrigation water via denitrification. J. Environ. Qual. 23:105-1 IO. Weil R.R., Weismiller R.A., Turner R.S. (1990) Nitrate contamination of groundwater under irrigated coastal plain soils. J. Environ. Qual. 19:441-448. Wilson M.. Lindow S.E. (1993) Release of recombinant microorganisms. Annu. Rev. Microbiol. 47:913-944. Wu Q.. Knowles R. (1994) Cellular regulation of nitrate uptake in denitrifying Flexibacrer canadensis. Can. J. Microbiol. 40576582. Wu Q.. Knowles R.. Niven D.F. (1994) 02regulation of denitrification in Flexibacter canadensis. Can. I. Microbiol. 40916-921. Ye R.W.. Averill B.A., Tiedje J.M. (1994) Denitrification: production and consumption of nitric oxide. Appl. Environ. Microbiol. 601053-1058. Yeomans J.C.. Bremner J.M., McCarty G.W. (1992) Denitrification capacity and denitrification potential of subsurface soils. Commun. Soil Sci. Plant Anal. 23:919-927. Zumft W.G., Dreush S., Lijchelt H., Cuypers B., Friedrich B., Schneider B. (1992) Derived amino acid sequences of rtflszgene (respiratory N20 reductase) from Alicaligenes eictrophus. Pseudonronos aerugimsa and Pseudomoms slutzeri reveal potential copper-binding residues. Eur. 1. Biochem. 208:3 1-40. Zumft W.G.. Korner H. (1997) Enzyme diversity and mosaic gene organization in denitrification. Antonie van Leeuweenhoek 71:43-58.

33 Tack for mig!

Forst och frlmst vi11 jag tacka alla de manniskor som pi nigot satt kanner sig ansvariga for att ha hjalpt mig nil hit. Ni skall ha all tack som glr att uppbringa!

Mannen som har haft ansvaret for min utbildning och i vissa fall bildning har varit Goran. Det finns minga superlativer jag skulle kunna anvanda for att beskriva Gorans egenskaper som handledare och van, men eftersom han inte tycker om sdnt tjafs si undviker jag det. Han har sina sidor, vissa som jag ratt snabbt larde mig forsti medan andra ar lika oforstieliga idag som de var for 6 ir sedan. Gorans forskarskola har varit tuff och ofta har jag tyckt att han varit orattvis och oforstlende, men han har nastan alltid lyckats motivera sina pistienden och It- garder pi ett vettigt satt. Vad man an tycker om Goran si ar han i grund och botten en trevlig prick! Tack for att du stitt ut med en cynisk och pessimistisk realist under alla dessa lr.

Marita gav mig ett brev efter ett mycket ostrategiskt samtal en tidig lordagsmorgon i tidernas begynnelse, diGCn pajat och allting verkade svart. Du skall veta att jag forsokt folja din rid, men som du si mycket val vet har det inte varit latt med Goran som handledare och ”meste medarbetare”. Jag vill ocksi passa pi att tacka for alla de trevliga och ofta smakliga stunder hemma hos Goran, Marita, Daniel och Liv.

I would like to thank David Richardson and Steven Spiro for introducing me to the mysteries of the molecular biology of denitrification and for sharing their vast knowledge. I had a great time during my stay at the UEA and hope to return some day.

Pelle Larsson har hjalpt mig med minga saker, allt frAn det byrikratiska tramset som av nlgon outgrundlig anledning miste goras till en del vetenskapliga frigor jag stalk. Han har dessutom forsett mig med en he1 del god litteratur (inte ryska klassiker, Sten!) som jag uppskattat oer- hort. Pelles alla rovarhistorier om den ldla konsten flugfiske har varit upplevelser som jag inte kommer att glomma i forsta taget.

Lennart Okla som villigt delat med sig av sina omfattande dator- och matematikkunskaper. Mycket av det jag nu kan om datorer, natverk, bytzoner och annat elande har jag Lennart at1 tacka for. Han har ocksi piggat upp det lite monotona avhandlingsarbetet pl sitt lite speciella satt som vi alla kanner.

Roland Lindqvist som hjalpte mig under de forsta kritiska &en. Utan din hjalp hade det defi- nitivt gitt it he1 ...

Anna Fossum, min rumskamrat under mlnga Hr. Vir, diskussioner, gral och skratt har alltid gjort det vart att leva. Jodi Anna, idag ar allt som vanligt, lika jivligt!!!

Christel Carlsson for alla trevliga pratstunder om allt mellan himmel och jord.

Anna Wallstedt, vlr egen emigrant, skall ha ett sort tack for att hon ar den hon ar. Du skall veta att jag jobbar pi en Kanada resa.

34 Och sedan har vi doktoranderna i ekotox gruppen som jag har stor respekt for. Kunnigare folk finns inte! De ekotoxare i min narmaste omgivning har varit,

Gudrun BremIe som stitt ut med alla mina fdgor om det mesta inom vetenskap och allskons andra omriden. Tack for all uppmuntran under avhandlingsarbetet. Det har varit guld vart!

,, , Olof Berglund som skall ha stort tack for all underhillning av mer musikalisk karaktar.

Goran Ewald som har gett rnig rid och tips om allt mijjligt.

Darius Sabaliunas, the prize winner, may you live long and prosper!

Annelie och Gunilla, vad hade vi gjort utan er! Tack for allt och dH menarjag allt!

Birgitta, min alskade, BomDI' 'IwwIj qaqaw jIH muSHa' SOH

Sist men inte minst, min mor som alltid knuffat mig framit de gilnger jag tappat tron pA hela skiten. Hon har hela tiden foljt mitt arbete med intresse och stallt mer eller mindre avancerade frigor om det jag pysslat med. Mina svar har inte alltid varit si latta att forsti, men jag har alltid forsokt forklara det svenska och inte pi "vetenskapitiska". Tack for allt ditt stod och knuffande.

Avslutningsvis vill jag bara saga att jag har lart mig mycket har pi Ekologen under &en som gitt. Jag kanner mig nu ratt val rustad for att tampas med mygel, svek, byrikrati och territori- ella beteenden.

Qapla'