Microbes Regulate Metal and Nutrient Cycling in Fe–Mn Concretions of the Gulf Finland

YEB

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Faculty of Biological and Environmental Sciences Department of Biosciences General Microbiology Doctoral Programme in Microbiology and Biotechnology University of Helsinki

Microbes regulate metal and nutrient cycling in Fe–Mn concretions of the Gulf of Finland

Pirjo Yli–Hemminki

Academic dissertation

To be presented for public examination with the permission of the Faculty of Biological and Environmental Sciences of the University of Helsinki in auditorium 1, Viikki Infocenter Korona (Viikinkaari 11, Helsinki) on November 13th at 12 o’clock noon.

Supervisors Doc Jouni Lehtoranta Marine Research Centre, Finnish Environment Institute

Doc Kirsten Jørgensen Marine Research Centre, Finnish Environment Institute

Thesis advisory Prof Kielo Haahtela (In memoriam) committee Department of Biosciences University of Helsinki, Finland

Doc Susanna Hietanen Department of Environmental Sciences University of Helsinki, Finland

Doc Petri Auvinen Institute of Biotechnology University of Helsinki, Finland

Prof Juha Karhu Department of Geosciences and Geography University of Helsinki, Finland

Reviewers Prof Marko Virta Department of Food and Environmental Sciences Univesity of Helsinki, Finland

Assoc Prof Sven Blomqvist Department of Ecology, Environmental and Plant Sciences Stockholm University, Sweden

Opponent Prof Lise Øvreås Department of Biology University of Bergen, Norway

Custos Prof Dennis Bamford Department of Biosciences University of Helsinki, Finland

Photo by the author: Fe–Mn concretions from the Gulf of Finland. ISBN 978-951-51-1623-9 (paperback), ISBN 978-951-51-1624-6 (PDF) ISSN 2342-5423 (paperback), ISSN 2342-5431 (PDF) Electronic version available at http://ethesis.helsinki.fi Hansaprint, Helsinki 2015 2

Abstract

Mineral concretions or nodules are found from the Oceans, lakes and in soils. Their element content has been studied well due to possible commercial use, but interest in their bacterial communities has risen due to environmental implications. Iron–manganese (Fe– Mn) concretions cover vast bottom areas in the Gulf of Bothnia and the Gulf of Finland (GoF). These mineral precipitates sequester several times higher amounts of Fe, Mn, phosphorus (P) and arsenic (As) than the surrounding sediment. Despite their large occurrence, the environmental significance of the concretion bottoms has been a somewhat understudied issue. The aim of the present study was to investigate the bacterial community structure, and possible microbial contributions to the formation and decay of concretions in the Baltic Sea. The further aim was to study how concretions respond to different environmental stresses, such as anoxia and crude–oil contamination, which the concretions may encounter in GoF bottoms. The methods used were determination of solid and dissolved Fe, Mn, P and As during microcosm incubations under oxidising or reducing conditions, also with crude oil and 14C–naphthalene added. Bacterial community structure was studied by cloning and sequencing taxonomic marker gene (partial 16S rRNA gene), and quantification of polycyclic aromatic hydrocarbon degradation (PAH–RHDα) gene copy number. Morphologically and taxonomically diverse bacteria colonise both the pitted surface and the porous interior of spherical concretions. Half of the population was affiliated to uncultured Proteobacteria, and one third was unclassified bacteria. Concretion bacteria populations appeared typical for this habitat. Bacteria may significantly affect the formation of the concretions in the GoF, because known Mn–oxidising bacteria were enriched in Mn2+–containing liquid and semi–solid media. The Fe2+–oxygen gradient favoured the enrichment of species which are known to reduce Fe and to degrade petroleum hydrocarbons. Concretions released Fe and Mn during anoxic conditions only if they were supplied with labile carbon source, indicating bacterial metal reduction. The dissolution of Mn was the highest, but the release of P and As followed Fe. The release rates (μmol m-2 d-1) from the concretions were within the range of the previously estimated fluxes out of the GoF sediment. Still, the concretions released only minor proportions (0.1–0.4%) of their total Fe, Mn, P and As content during a prolonged anoxic period. Concretions and sediment had a very similar capacity to remove petroleum compounds and mineralise naphthalene under oxic as well as anoxic conditions, and over one half of the added crude oil disappeared from the microcosms. Copy numbers of PAH– degradation genes increased, indicating biological degradation potential of PAHs by the concretion bacterial community. Both concretions and sediment had rich and clearly different bacterial communities prior to and also past the exposure to crude oil. Only 9% of the OTUs were shared between the initial concretions and the sediment. Concretion bacterial sequences were affiliated to bacterial groups previously found from concretions and metal rich environments (ecotypes) even after the crude–oil exposure, whereas 3 sediment bacterial sequences were similar to those originating from sediments and oil– contaminated sites.

Tiivistelmä

Konkretiot (myös nk. nodulit) ovat mineraalisaostumia, joita esiintyy merien ja järvien pohjilla sekä maassa. Konkretioiden alkuainekoostumusta on tutkittu paljon niistä mahdollisesti saatavan raaka–aineen takia. Kiinnostus konkretioiden mikrobiyhteisöjä kohtaan on myös kasvanut arvioitaessa konkretioiden merkitystä ympäristössään. Rauta– mangaanikonkretioita (Fe–Mn) löytyy laajoilta alueilta Pohjan– ja Suomenlahden pohjilta, joissa niihin on saostunut huomattavasti suurempia määriä rautaa, mangaania, fosforia (P) ja arseenia (As) kuin ympäröivään sedimenttiin. Laajuudestaan huolimatta konkretiopohjien merkitystä ympäristölleen on tutkittu vähän. Tämän työn tarkoitus oli tutkia konkretioiden bakteeriyhteisöjen rakennetta ja mikrobien vaikutusta konkretioiden muodostumiseen ja hajoamiseen Itämeressä. Tutkittiin myös kuinka erilaiset ympäristön stressitekijät, kuten hapettomuus ja raakaöljyllä saastuminen, voivat vaikuttaa konkretiopohjiin Suomenlahdella. Konkretioista määritettiin niiden Fe, Mn, P ja As pitoisuudet. Mitattiin myös kuinka paljon ja miten nopeasti määritetyt aineet liukenivat konkretioista eri koeolosuhteissa. Konkretionäytteitä inkuboitiin mikrokosmoksissa hapellisissa ja hapettomissa olosuhteissa, myös raakaöljyn ja radioisotoopilla leimatun 14C–naftaliinin kanssa. Bakteeriyhteisöjä tutkittiin kloonaamalla ja sekvensoimalla bakteereille spesifejä 16S rRNA geenejä, joiden vastaavuutta verrattiin sekvensseihin kansainvälisissä tietokannoissa. Bakteerien geneettistä potentiaalia hajottaa polyaromaattisia hiilivetyjä (PAH) tutkittiin mittaamalla PAH–RHDα geenin lukumäärää kvantitatiivisella PCR–menetelmällä. Monimuotoinen bakteerilajisto elää pallomaisten konkretioiden röpelöisellä pinnalla ja huokoisessa sisuksessa. Puolet lajistosta voitiin luokitella kuuluvaksi Proteobakteereihin mutta kolmasosa bakteerilajeista oli tuntemattomia. Konkretioiden bakteeriyhteisö ilmeni olevan tälle habitaatille tyypillinen (ekotyyppi). Lähimmät vastaavuudet sekvenssitietokannassa olivat peräisin konkretioista tai muista metalleja sisältävistä ympäristöistä. Bakteerit voivat merkittävästi vaikuttaa konkretioiden muodostumiseen Suomenlahdella. Koeolosuhteissa konkretioista peräisin olevat bakteerit hapettivat mangaania ja osallistuivat raudan hapettumiseen. Hapettavissa olosuhteissa rikastuneisiin bakteereihin kuului muun muassa tunnettuja mangaania hapettavia lajeja ja rautaa pelkistäviä lajeja. Hapettomissa olosuhteissa konkretioista vapautui rautaa ja mangaania merkittäviä määriä vain silloin kun metalleja pelkistäville konkretiobakteereille oli tarjolla lisätty hiilen lähde. Mangaania liukeni eniten mutta fosforin ja arseenin vapautuminen oli riippuvaista raudan liukenemisesta konkretioista. Metallien liukenemisnopeudet (μmol m-2 d-1) konkretioista olivat samaa suuruusluokkaa kuin aikaisemmin arvioidut nopeudet

Suomenlahden sedimenteistä. Siitä huolimatta, ainoastaan alle prosentti (0.1–0.4%) konkretioiden sisältämästä raudasta, mangaanista, fosforista ja arseenista vapautui usean hapettoman koeviikon aikana. Konkretiot saattavat siis pidättää näitä aineita paremmin kuin ympäröivä sedimentti kausittaisten hapettomien jaksojen ajan Suomenlahdella. Konkretioilla ja niitä ympäröivillä sedimenteillä oli samantasoinen kyky poistaa raakaöljyä ja hajottaa naftaleenia hapellisissa ja hapettomissa olosuhteissa. Vähintään puolet lisätystä raakaöljystä katosi mikrokosmoksista biohajoamisen ja osin kemiallisen hajoamisen/ sitoutumisen takia. PAH–RHDα geenin lukumäärää kasvoi, mikä viittasi konkretiobakteerien mahdolliseen kykyyn hajottaa PAH–yhdisteitä. Konkretioiden ja niitä ympäröivien sedimenttien bakteeriyhteisöt poikkesivat toisistaan aluksi ja jopa raakaöljyaltistuksen jälkeen. Vain noin 9% taksoneista oli konkretioissa ja sedimentissä samoja. Raakaöljylle altistaminen ei vähentänyt bakteeriyhteisöjen monimuotoisuutta tai johtanut tunnettujen öljynhajottajalajien rikastumiseen. Sedimenttiä sisältävissä mikrokosmoksissa bakteerisekvenssien lähimmät vastaavuudet löytyivät sedimenteistä ja öljyllä saastuneista paikoista. Konkretioiden ominaisuudet ovat suunnanneet bakteeriyhteisöjen muodostumista tälle mikrohabitaatille tyypilliseksi.

Table of contents

Abstract ...... 3 Tiivistelmä ...... 4 List of original publications ...... 8 Abbreviations ...... 9 1. Introduction ...... 10 1.1. Fe–Mn concretions in the Baltic Sea ...... 11 1.2. Microbe–metal interactions ...... 13 1.2.1. Fe–oxidising metabolism ...... 13 1.2.2. Fe–oxidising bacteria in their natural habitats ...... 14 1.2.3. Mn–oxidising metabolism ...... 15 1.2.4. Mn–oxidising bacteria in their natural habitats ...... 15 1.2.5. Fe– and Mn–reducing metabolism ...... 16 1.2.6. Fe and Mn reduction in marine sediments ...... 17 1.2.7. Fe– and Mn–reducing bacteria in marine sediments ...... 18 1.3. Environmental implications ...... 18 1.3.1. Eutrophication in the Gulf of Finland ...... 18 1.3.2. Bioremediation of harmful metals ...... 19 1.3.3. Oil–degradation potential in the sediments ...... 20 1.4. Aims of the study ...... 22 2. Materials and methods ...... 22 2.1 Study sites and sampling procedure ...... 22 2.2. Experimental set ups and methods ...... 23 3. Results ...... 27 3.1 Physical and geochemical characteristics of the GoF concretions ...... 27 3.2 Fe and Mn –oxidising and –reducing processes in the GoF concretions ...... 28 3.3 Disappearance of crude oil and mineralisation of 14C naphthalene ...... 31 3.4 Bacterial communities of the GoF concretions ...... 33 3.5 Mn– and Fe–oxidising bacteria enriched from the concretions ...... 35 3.6 Bacterial community responses to crude–oil exposure ...... 35 4 Discussion ...... 36 4.1. Concretions as microbial habitats in the GoF ...... 36 4.2. Bacterial community structure of the Fe–Mn concretions in the GoF ...... 37 4.3. Bacterial Fe and Mn oxidation contribute to the formation of the GoF concretions 38

4.4. Microbial Mn and Fe –reduction control P and As release from the concretions ... 39 4.4.1. Effects of temperature on Mn and Fe –reduction and P and As release from the concretions ...... 40 4.4.2. Release of Fe, Mn, P and As vs. sediment effluxes ...... 41 4.5. Degradation potential of crude oil and PAHs by Fe–Mn concretions in the GoF ... 42 4.5.1. Fe and Mn reduction under the crude oil exposure ...... 43 4.5.2 Abiotic and biotic disappearance of crude oil and PAHs ...... 43 4.5.3. Concretion bacterial community responses to petroleum hydrocarbons ...... 44 Conclusions ...... 45 Future prospects ...... 46 Acknowledgements ...... 47 References ...... 48

List of original publications

This thesis is based on the following publications and manuscripts, which are referred to in the text by their Roman numerals:

I Yli–Hemminki P, Jørgensen KS and Lehtoranta J (2014) Iron–manganese concretions sustaining microbial life in the Baltic Sea: The structure of the bacterial community and enrichments in metal–oxidizing conditions. Geomicrobiology Journal 31:263–275.

PYH was responsible for designing the experiment, conducted the experimental work and processed the data, and was the corresponding author.

II Yli–Hemminki P, Sara–aho T, Jørgensen KS and Lehtoranta J (2015) Iron– manganese concretions contribute to benthic release of phosphorus and arsenic in anoxic conditions in the Baltic Sea. Submitted manuscript

PYH was responsible for designing the experiment, conducted the experimental work and processed the data, and was the corresponding author.

III Reunamo A, Yli–Hemminki P, Nuutinen J, Lehtoranta J and Jørgensen KS (2015) Degradation of crude oil and PAHs in the iron–manganese concretions and sediment bottoms in the northern Baltic Sea. Submitted manuscript

PYH designed the experiments, conducted the experimental work and processed the data with AR and shared the authorship with AR. This manuscript has also been published as a part of the thesis of AR.

The published articles were reprinted with the permission from the copyright holders.

Abbreviations

16S rRNA 16S subunit ribosomal ribonucleic acid DAPI 4’,6’–diamidino –2–phenylindole DGGE denaturing gradient gel electrophoresis ENA European Nucleotide Archive GC–FID gas chromatograph – flame ionization detector GN Gram negative GoF Gulf of Finland GP Gram positive ICP–OES inductively coupled plasma – optical emission spectrophotometer ICP–MS inductively coupled plasma – mass spectrometer LOI loss on ignition ORF open reading frame OTU operational taxonomical unit PAH polycyclic aromatic hydrocarbon PAH–RHDα PAH ring–hydroxylating dioxygenase PCR polymerase chain reaction RFLP restriction fragment length polymorphism

1. Introduction

Microorganisms (here prokaryotes: Archaea and Bacteria) are ubiquitous and they are present also during conditions and in environments which seem hostile. Bacteria are colonisers of solidified lava, i.e. basalts, of the ocean crust (Edwards et al. 2005, Santelli et al. 2008), and of rocks where they may induce precipitation of minerals (reviewed by Raab and Feldmann 2003, Vasconcelos and McKenzie 2009). Microbes are by definition small but because of their countless number their influence on their surroundings is enormous (Whitman et al. 1998). Bacteria are involved at every trophic level, from primary production to decomposition because they have got diverse and versatile metabolisms to process organic and inorganic compounds to meet their nutritional and energetic needs (Fierer and Lennon 2011). Different species commonly live closely, also in biofilms (reviewed by Hall–Stoodley et al. 2004), where bacteria use available oxidants in the order - 2+ starting from those yielding the greatest free energy (O2 > NO3 > MnO2 > FeOOH > SO4

> CO2), and where the metabolic waste from one species is often substrate for another. Many bacteria are able to couple metal (Fe, Mn, Cr, U) reduction to oxidation of energy sources (Nealson et al. 2002). Reduced metals may be oxidised again, which leads to redox cycle of the metals (Roden et al. 2004), and this affects also the mobility of associated elements (P, As). Other life forms are crucially dependent on this reworking of compounds and recycling of elements throughout the biosphere. Human impact on the environment has had undesired consequences, such as eutrophication of water systems and pollution of sediments. Here we again rely on microbial treatment as remedy. Bacteria control the compositions of soil, sediments and water by degrading harmful compounds and by transforming and precipitating harmful elements. We need more knowledge about these processes in order to better manage the environmental problems. Microbial ecologists are interested in i.a. activity and diversity of bacterial populations in the environment. Geomicrobiologists study microbe–mineral interactions and their environmental implications. State of the environment, e.g., sediment, can be monitored by measuring essential parameters (T °, pH, redox potential, water content, concentrations of concerned substances and elements) over time. Microbial processes, e.g., Fe oxidation–reduction, can be monitored by bringing intact sediment cores to a laboratory, where they can be incubated in controlled conditions while samples are taken for Fe analyses. Subsamples from sediment can also be treated in carefully designed micro– or macroscale incubations, which simulate natural conditions. After the initial parameters have been measured in such microcosms, bacterial metabolites and changes in the concentrations of the measured substances and degradation products can be traced. Interesting substrate (e.g. compounds of petroleum hydrocarbons; reviewed by Widdel and Rabus 2001) can be labelled with its stable or radioactive isotope which is used to trace its degradation pathway or mineralisation rate and metabolic utilisation by bacteria. Nitrate 15 3- 35 2+ (stable NO ) and sulphate (radiolabeled SO4 ) reduction rates are commonly quantified by isotope pairing methods (Jørgensen 1978, Nielsen 1992, Steingruber et al. 2001). It is possible to combine isotope techniques to DNA techniques in order to trace 10 which bacteria use the added labelled substrate (e.g., 13C–labelled carbon source) as building blocks of their DNA (Radajewski et al. 2000). Connections between environmental characteristics, bacterial community structure and genetic potential of the microbial communities can be studied by combining physicochemical measurements, metagenomics approach and detection of certain bacterial metabolic marker genes e.g. from a stratified water body sampling site (Llorens–Mare`s et al. 2015). Cultivation, isolation and characterisation of bacterial species from different environments are a never–ending task of microbiologists. It is time taking and challenging, and only a small fraction of bacteria may be studied in laboratory conditions (Epstein 2013). Nevertheless, the species diversity can be evaluated by studying molecular data, most typically bacterial 16S rRNA gene sequences obtained from the environmental sample (reviewed by Hugenholtz et al. 1998). The 16S rRNA sequences are compared to each other, and defined as operational species, i.e., operational taxonomic units (OTU), if the sequences are more than 97% different. The growing amount of sequence data is collected to international data banks, such as European Nucleotide Archive (ENA), which has got daily data exchange with GenBank and DNA Data Bank of Japan (Leinonen et al. 2010) and Ribosomal Database Project (Wang et al. 2007), where the data is available for researchers. Also, microscopy techniques can be combined to DNA techniques for metabolic studies. Fluorescence in situ hybridization (FISH) methods utilise fluorescent DNA probes of known bacterial sequences (Amann et al. 1995, Moter and Göbel 2000). The probes can be targeted to 16S rRNA genes or to functional genes in the sample, and it is possible to visualise the targeted bacteria by microscope. Target genes and their expression (mRNA) can be quantified by qPCR (Heid et al. 1996, Rasmussen et al. 1998), which is a sensitive method to detect even low number of functional genes from the sample. The development of the quantitation capacity of computers and the increase in DNA–sequencing speed has enabled sequencing of whole genomes, transcripted genes (total mRNA) and total bacterial communities, and has brought sequencing techniques to the next level referred as Next Generation Sequencing (NGS). The new techniques allow better representation of sample diversity than before (Shokralla et al. 2012). Full published genomes of 158 Fe–oxidising and –reducing bacteria and archaea were analysed to reveal the open reading frames (ORF) of genes within the genomes (Alterman 2014). These ORFs were compared to each other by functional genome distribution analysis to see genome to genome similarities. The analysis visualised that Fe–oxidising and –reducing microbes mostly fall in separate clusters (Alterman 2014).

1.1. Fe–Mn concretions in the Baltic Sea

Large deposits of Fe–Mn concretions (also called as marine nodules) are situated on the surface of silty–clayey mud bottoms in the Gulf of Bothnia (GoB) and the Gulf of Finland (GoF), which are parts of the world’s second largest brackish–water basin, the Baltic Sea.

Concretions are mineral precipitates which vary in shape being flat crusts, discoid or spherical, from millimetres to several centimetres in diameter (Ingri 1985). Concretions were first discovered from the Oceans, but they are also found at lake bottoms and soil. Dense deposits of concretions cover approximately 10% of the bottom areas of the GoB and the GoF (Winterhalter 1966, Ingri 1985). The density can be as high as 17 kg m-2 in the Gulf of Riga (reviewed by Baturin and Dubinchuk 2009), 15 kg m-2 in the GoB (Ingri 1985) and 24 kg m-2 in the eastern GoF (Glasby et al. 1997). During their formation period concretions capture large amounts of Fe, Mn, P and As as well as other trace elements, and their concentrations are higher than in the surrounding sediment (Ingri 1985, Baturin 2009, Baturin and Dubinchuk 2009). In the Baltic Sea concretions, Fe usually appears as goethite and amorphous Fe hydroxide phases (Ingri 1985). Arsenic and P are closely related to Fe in the concretions (Ingri 1985, Marcus et al. 2004). Birnessite and rare manganosite phases of Mn have been reported from Baltic Sea concretions (Zhang et al. 2002). The concretions may attract commercial interest as Mn ore (Ingri 1985, Rona 2003, Anufriev et al. 2005). Formation–rate estimate for the concretions is 14 μm a-1 (Grigoriev et al. 2013), which is considerably higher than those of oceanic concretions, 0.01 μm a-1 (Glasby et al. 1982). The short geological history of the Baltic Sea reveals that concretions are rather young (maximum 3000 years; Ingri 1985), their mean age being less than thousand years, between 670–850 years (Grigoriev et al. 2013). Content of the concretions may reflect changes in their formation environment as well as anthropogenic load of harmful elements to the surrounding sediment (Ingri 1985, Leivuori and Niemistö 1995, Emelyanov and Kravtsov 2007, Baturin and Dubinchuk 2009). Microorganisms are involved in the transformation of rocks and minerals at the seafloor (Edwards et al. 2005). Precipitation mechanisms of Fe and Mn oxides (hydroxides, oxyhydroxides or oxides) on the concretions have traditionally been explained by plain equilibrium geochemistry (Baturin and Dubinchuk 2009). The role of specific bacteria able to oxidize and accumulate Mn has also been proposed, but specific microbiological studies of concretions are rare (samples from Atlantic Ocean by Ehrlich 1963, from western Baltic Sea by Ghiorse and Hirsch 1982). Microstructures of bacteria and possibly bacterially formed manganosite (MnO) have been identified in the concretions of the GoF (Zhang et al. 2002), and in biofilms attached to eastern Pacific Ocean nodules (Wang et al. 2009), and lipid biomarkers characteristic to bacteria in salt marsh concretions (Duan et al. 1996) have been observed, encouraging to further investigate the role of bacteria in the formation of concretions. Development of DNA– sequencing methods and databases has enabled studies of bacterial community structure of concretions regardless of difficulties in culturing the bacteria. Study of concretions in the Pacific Ocean (Wang et al. 2009) showed that some of the cloned sequences of bacterial 16S rRNA genes had a close relationship with sequences obtained from soil concretions (He et al. 2008, Zhang et al. 2008), which in turn had diverse bacterial communities similar to ones in metal–rich environments, and with the presence of some strains related to the known Mn–oxidising bacteria from the aquatic environment. 12

1.2. Microbe–metal interactions

Reactive, negatively charged sites on bacterial surfaces commonly cause aggregation of metal–oxide particles on bacteria (Mullen et al. 1989, Templeton et al. 2005). Bacteria may thus act as nucleating agents precipitating Fe and Mn oxides that are not enzymatically formed. This encrusts the microbial cells in the oxides. Metals can be divided into neutral elements, essential nutrients and elements toxic for organisms. Dissimilatory Fe and Mn oxidation and reduction refer to processes by which bacteria produce energy.

1.2.1. Fe–oxidising metabolism

Bacteria are able to oxidise reduced species of Fe2+ and Mn2+/3+ and thereby precipitate and accumulate metal oxides, for instance in sediments (Emerson 2004, Tebo and Villalobos 2005). In environments rich in ferrous iron (Fe2+), bacteria can gain energy for growth from dissimilatory Fe oxidation as follows: 2+ + 3+ 4Fe + O2 + 4H –> 4Fe + 2H2O Energy yield derives from the great difference between the electrode potentials (ΔE°) of electron donating Fe2+ and electron accepting oxygen. Among the first Fe–bacteria described were acidophilic Acidithiobacillus ferrooxidans (previously Thiobacillus ferrooxidans; Temple and Colmer 1951) and Leptospirillum ferrooxidans. They grow e.g. in mine–waste disposal sites where pH 2 or below hinders abiotic oxidation of Fe2+. Bacteria may use energy from the oxidation reaction to synthesise organic compounds (autotrophy), or they may require additional carbon source (mixotrophy) for their growth. Fe oxidation among heterotrophic bacteria may allow bacteria to adapt to changing environmental conditions (Edwards et al. 2003a). Further sign of versatility is that, e.g. Rhodoferax ferrireducens T118 and Geobacter bremensis R1 are able to either oxidise or reduce Fe, depending on environmental conditions (Alterman 2014). Phylogenetically diverse purple sulphur, non–sulphur and green–sulphur bacteria are able to photoautotrophic oxidation of Fe2+ in anoxic conditions (reviewed by Weber et al. 2006 and Bird et al. 2011). Furthermore, bacteria have been observed to oxidise Fe2+ under - anoxic conditions using nitrate (NO3 ) as an electron acceptor as follows: 2+ - 3+ - 10Fe + 2NO3 + 6H2O –> 10Fe + N2 + 12OH (Straub et al. 1996) These autotrophic or mixotrophic bacteria have been isolated from sediment samples from freshwater and brackish water environments (Straub et al. 1996, Straub and Buchholz– Cleven 1998, Hauck et al. 2001, reviewed by Weber et al. 2006). There are members of Proteobacteria, and also known Fe reducer, Geobacter metallireducens, which oxidise Fe - while reducing NO3 (Finneran et al. 2002, reviewed by Hedrich et al. 2011). Microaerobic, autotrophic bacteria, Sideroxydans lithotrophicus and Gallionella capsiferriformans, have been identified to use an extracellular decaheme c–type 13 cytochrome to oxidise Fe in fresh water environments (Liu et al. 2012, Emerson et al. 2013). To date, a number of bacterial (54) and archaeal (13) genomes of Fe oxidisers have been published (reviewed by Weber et al. 2006 and Alterman 2014). Unfortunately, bacteria do not share a gene which encodes an enzyme by which they oxidise Fe (reviewed by Weber et al. 2006, Singer et al. 2013), and this further complicates the development of so called universal DNA–primer for Fe–oxidising bacteria.

1.2.2. Fe–oxidising bacteria in their natural habitats

Phylogenetic studies of the bacterial populations associated with different neutral Fe–rich environments have revealed very diverse communities (e.g., Bruun et al. 2010 and references therein). Due to the high diversity, the contribution of many different species to Fe oxidation remains unresolved. Culture–based experiments may enrich new Fe– oxidising species and help to connect the process to the species. Cultivated microaerobic Fe–oxidising bacteria growing at circumneutral pH have often been recognised as Proteobacteria: Betaproteobacteria in fresh water, and members of a new candidatus class Zetaproteobacteria in marine environments (reviewed by Emerson et al. 2010 and Hedrich et al. 2011). In near–neutral conditions (pH 6–8), bacteria compete with chemical oxidation of Fe2+, and in nature they have to position themselves at low oxygen concentrations (< 1mg L-1) (Konhauser et al. 2011). Microaerophilic bacteria are found from oxic–anoxic interfaces rich in Fe (reviewed by Emerson et al. 2010). Bacteria are associated with Fe oxides at fresh–water seepages where they actively mediated even 80% of the Fe oxidation (Emerson and Revsbech 1994a, Emerson and Revsbech 1994b, Bruun et al. 2010). Neutrophilic Fe–oxidising bacteria, also other than classical Gallionella ferruginea (Hanert 1992) and Leptothrix ochracea (Corstjens et al. 1992), have been sampled and cultivated from drains filled with Fe–rich groundwater (Emerson and Moyer 1997), and around the roots of wetland plants (Emerson et al. 1999) from where new species called Ferritrophicum radicicola and Sideroxydans paludicola were characterized (Weiss et al. 2007). Environmental geochemical characteristics such as reduced and oxidised metals in marine basalts have encouraged the search for novel neutrophilic Fe–oxidising bacteria (Mason et al. 2009, reviewed by Emerson et al. 2010). It has been inferred that basaltic Mn, S, and Fe phases may supply energy for chemolithoautotrophic growth of microorganisms in deep–sea environments (Santelli et al. 2008) because basalts are likely to be the largest reservoir of reduced Fe in the Earth’s crust (Edwards et al. 2004). Fe oxidation may then provide energy for significant carbon assimilation and bacterial biomass production in the crust (Edwards et al. 2004). Bacteria indeed colonise ocean crust sites rich in Fe–oxide accumulations, and culture–dependent studies have shown that they are microaerophilic, chemolithoautotrophic Fe–oxidising bacteria (Edwards et al. 2003a). These psychrophilic bacteria grow at ambient temperature 3–10 °C. Phylogenetic analysis

14 of these strains showed similarity to uncultured Alpharoteobacteria and Marinobacter sp. (Gammaproteobacteria), which not earlier have been known from Fe–oxidation context (Edwards et al. 2003b). Fe–oxidising bacteria are also abundant at Loihi hydrothermal vent sites near Hawai, where microbial mats are found encrusted with Fe oxides (Emerson and Moyer 2002). Cultivation and characterisation of these enriched bacteria has led to the discovery of a new chemolithotrophic Fe–oxidising species Mariprofundus ferrooxydans in a new candidate class of (ζ–) Zetaproteobacteria (Emerson et al. 2007, Singer et al. 2013). Also Fe–oxidising archaeon, classified to Ferroplasmaceae family, has been isolated from bioleaching pilot plant (Golyshina et al. 2000).

1.2.3. Mn–oxidising metabolism

Mn2+ is not as readily oxidised chemically as Fe2+ is, and bacterially catalysed Mn2+ oxidation is responsible for the bulk of the Mn–oxide formation in the sediments

(Thamdrup et al. 1994). Biogenic Mn oxides are best identified as δ–MnO2–like poorly crystalline phyllomanganates (Tebo et al. 2004). Dissolved oxygen may be the only oxidant for the biologically catalysed Mn2+ oxidation (Schippers et al. 2005, Clement et al. 2009) in contrast to documented anaerobic Fe oxidation with nitrate in anoxic environment (Straub et al. 1996). Single bacterium can oxidise Mn2+ straight to Mn4+ using e.g. multicopper oxidase–like enzyme (Tebo et al. 2004): 2+ + 3+ 2Mn + 0.5O2 + 2H –> 2Mn + H2O 3+ + 2Mn + 0.5O2 + 3H2O –> 2MnO2 + 6H The contribution of bacteria to Mn oxidation is not as controversial as to Fe oxidation, since biological Mn2+ oxidation is much faster than abiotic reaction (Diem and Stumm 1984, Tebo and Emerson 1986, Tebo et al. 2004). Still, the reason why bacteria oxidise Mn is not that clear (Tebo et al. 2005). Only some of the known Mn–oxidising bacteria may use the reaction for energy production (e.g., Pseudomonas and Bacillus species), while other species may oxidise Mn to protect the cell from the toxicity of reactive oxygen species (Lactobacillus plantarum; Archibald and Fridovich 1981), other toxic metals or UV radiation (Deinococcus radiodurans) (Tebo et al. 2005 and references therein).

1.2.4. Mn–oxidising bacteria in their natural habitats

Metal–oxidising biofilms form ferromanganese crusts in cave systems of Tennessee River system (Carmichael et al. 2013). Mn–oxidising species, such as Leptothrix, Pseudomonas, Flavobacterium and Janthinobacterium,where cultured and identified therein from otherwise diverse bacterial community, and many other recorded OTUs were related to cloned sequences isolated from metal–rich environments. Potential synergistic interactions 2+ 2+ may occur between CO2–fixing Fe –oxidizing bacteria and heterotrophic Mn –oxidizing 15 bacteria in biofilms during the generation of ferromanganese crusts on basalts (Templeton et al. 2005). Several new heterotrophic, Mn–oxidising Alpha– and Gammaproteobacteria such as Sulfitobacter, Methylarcula and Pseudoalteromonas and additionally Microbulbifer, Alteromonas, Marinobacter and Halomonas have been isolated from submarine basalts of Loihi, Hawai. Also, Fe–Mn concretions have been recognised as environments likely to harbour Mn–oxidising bacteria (Tebo et al. 2005). Some of the early experiments of biological Mn oxidation have been carried out using bacteria or cell–free extracts of the bacteria from soil concretions (Pseudomonas sp. and Citrobacter freundii; Douka 1980) and from Atlantic marine concretions (Arthrobacter; Ehrlich 1968). Within only weeks, bacteria can foul and form Mn micronodules on the metal surfaces of heat exchangers that use Baltic Sea water (Kuosmanen et al. 2005). Bacterial Fe and Mn reduction produces soluble species of the metals which diffuse from an anoxic sediment zone to oxic one to be readily oxidised by bacteria (Bruun et al. 2010).

1.2.5. Fe– and Mn–reducing metabolism

Concretions do not withstand high sedimentation rates in the GoB and the GoF (Ingri 1985) and they start to decay if they get buried under sediment layers greater than 50–100 mm (Zhamoida et al. 2007). The decay may be caused by chemical or bacterial reduction and dissolution of Fe and Mn, because Fe and Mn oxides in the concretions may be bioavailable. Reduced Fe and Mn are soluble. In the absence of oxygen many bacteria can couple their growth to Fe and Mn reduction (Myers and Nealson 1988, reviewed by Lovley 1991). The same bacteria are usually able to reduce both Fe and Mn (Lovley and Phillips 1988). Fe3+ commonly follows Mn4+ in the sequence of electron acceptors based on the free energy yield. Bacteria may also control the toxicity of the metals and metalloids by reducing them. Fe and Mn serve as oxidants of simple carbon compounds: - - 2+ - CH3COO + 8FeOOH + 3H2O -> 2HCO3 + 8Fe + 15OH - - 2+ - CH3COO + 4MnO2 + 3H2O -> 2HCO3 + 4Mn + 7OH The crystal structure of Fe oxides effect on how accessible they are for Fe–reducing bacteria (Roden and Zachara 1996), and thus ferrihydrite is favoured due to its poorly ordered crystallinity (reviewed by Straub et al. 2001). Fe minerals produced by Fe– oxidising bacteria are good electron acceptors for Fe–reducing bacteria (Emerson and Revsbech 1994a). Bacteria have to possess the cellular mechanism by which they transfer electrons to precipitated Fe3+. The most extensively studied Fe–reducing bacteria are probably Shewanella and Geobacter species and their electron–transfer mechanisms (reviewed by Bird et al. 2011). Bacteria may transfer electrons to Fe3+ minerals directly from outer membrane c–type cytochromes, indirectly via electron shuttles (humic substances or cellular secretions), or indirectly via the solubilisation of Fe3+ by organic

16 chelators (reviewed by Straub et al. 2001, Kappler et al. 2004, reviewed by Bird et al. 2011). Shewanella and Geobacter have also been shown to conduct electrons through long pili–like appendages called “nanowires” (Reguera et al. 2005). Network of bacterial nanowires may connect different locations in sediments (Nielsen et al. 2010, Pfeffer et al. 2012) and soil (Ball 2007), and allow bacteria to transfer electrons even to distant electron acceptors. Since the most extensively studied Fe– and Mn–reducing bacteria are able to transfer electrons to solid cathode, the bacteria can even be harnessed to produce electricity from aquatic sediments and waste organic matter (reviewed by Lovley 2004 and Lovley 2006). Bacterial dissimilatory Fe3+ and Mn4+ reduction can be coupled to oxidation of organic compounds or hydrogen. Simple carbon compounds (e.g., acetate and lactate) produced by fermenting bacteria can be oxidised by Fe and Mn–reducing bacteria (Vandieken et al. 2013), along with several other substrates (e.g., glucose, pyruvate, volatile fatty acids) that have been tested (reviewed by Lovley 1991 and Lovley 2004). Geobacter oxidises also harmful compounds, such as chlorinated compounds, benzene, phenol and toluene in anoxic conditions (reviewed by Lovley 1991, Lovley and Anderson 2000).

1.2.6. Fe and Mn reduction in marine sediments

Complex bacterial community competes over the same substrates in sedimentary environments. Usually, bacteria which use more energetically favourable electron 2+ acceptors prevail. Bacterial sulphate (SO4 ) reduction has been argued to be the most important anaerobic respiration pathway in marine environment (reviewed by Straub et al. 2001, Vandieken et al. 2006). Fe– and Mn–reducing bacteria can though outcompete 2+ SO4 –reducing bacteria and methanogens for electron donor if bioavailable Fe or Mn– oxide concentrations are relatively high in the sediment (Lovley and Phillips 1987, Canfield et al. 1993, Thamdrup et al. 2000, Jensen et al. 2003, Roden and Wetzel 2003). 3+ Amorphous Fe oxyhydroxide and MnO2 have got large available surface areas which are easily reduced by bacteria (Jensen et al. 2003, Spiro et al. 2008). Metal–reducing bacteria are dependent on the supply of Fe and Mn oxides, and the bacteria have to position themselves at the anoxic–oxic interface zone where the recycling of metals is possible by introduction of oxygen caused by bioturbation of the sediment (Canfield et al. 1993). Therefore, upper sediment layers (1–10 cm) may support Mn and Fe reduction whereas 2+ SO4 reduction dominates in the deeper layers (Canfield et al. 1993, Thamdrup et al. 2000, Vandieken et al. 2013). Fe or Mn reduction has been calculated to account for 20–78% of total anaerobic C oxidation in the sediments of eastern Skagerrag, (Canfield et al. 1993), and 60–75% in the northern and eastern Kattegat, Denmark (Jensen et al. 2003), and 13% in cold fjord sediment on the western Svalbard (Vandieken et al. 2006). Similar range (25–73%) of Mn reduction coupled to C oxidation has been calculated also in Black sea sediments with high

Mn concentration (Thamdrup et al. 2000), and (50%) in Mn–rich marine sediment from Gullmarsfjorden, Sweden (Vandieken et al. 2014). Bacteria may be selective over electron acceptors as discovered in Black Sea shelf sediment. There bacteria reduced Mn oxide much more efficiently than ferrihydrite (Thamdrup et al. 2000). Similar calculations have not been made for the gulfs of Bothnia or Finland.

1.2.7. Fe– and Mn–reducing bacteria in marine sediments

Many known isolated Fe reducers belong to Proteobacteria, and also the roles of Acidobacteria, Actinobacteria, Deferribacteres, Firmicutes and Archaeon have been demonstrated (reviewed by Lovley 2004, Weber et al. 2006 and Alterman 2014). Not all the known microorganisms conserve energy from this reaction though. Even if species of Geobacter (Deltaproteobacteria) and Shewanella (Gammaproteobacteria) are found from sedimentary environments and soil (Coates et al. 1996b, Lonergan et al. 1996, Cummings et al. 2003, Vandieken et al. 2006), the composition of Fe and especially Mn–reducing 2+ communities in marine sediments are still poorly known. Some SO4 –reducers are found to reduce Fe3+ as well (Coleman et al. 1993). Genera Desulfuromonas, Desulfuromusa and Desulfovibrio were identified large Fe reducers in cold fjord sediments of Svalbard (Vandieken et al. 2006). Many of the bacteria originally found Fe reducers are usually able to reduce Mn as well, but in nature the species composition responsible for these reactions may differ greatly (Vandieken et al. 2012). The Mn–reducing bacterium Alteromonas putrefaciens strain MR–I was isolated from the anaerobic lake sediments (Myers and Nealson 1988). Bacteria that specialise in Mn reduction were identified as Desulfuromonas, Pelobacter and Arcobacter type species and Shewanella in sediment rich in Mn oxides in Black Sea shelf (Thamdrup et al. 2000). Acetate–oxidising Mn–reducing bacteria were similar, Arcobacter, Colwellia and Oceanospirillaceae, even in geographically distinct Mn–rich sediments in Gullmarsfjorden (Sweden), Skagerrak (Norway) and Ulleung Basin (South Korea) (Vandieken et al. 2012). Deep subsurface environments are isolated, and harbour distinct bacterial community involved in Fe reduction (Dong et al. 2014). Species related to Vulcanibacillus and Orenia were enriched, and are possibly indigenous to the Mt. Simon Sandstone, U.S.A., at 2 km depth.

1.3. Environmental implications

1.3.1. Eutrophication in the Gulf of Finland

Eutrophication is a worldwide problem in coastal areas where riverine runoff supplies the waters with excess nutrients (e.g., fertilizers) which in turn feed the mass growth of algae. Enhanced primary production leads to accumulation of organic matter, and mineralization of this matter leads to increased oxygen consumption in bottom waters (reviewed by Diaz 18 and Rosenberg 2008). Ecosystem responses to hypoxia (< 2 mL O2/ L) are mortality or get away of benthic organisms, leaving anaerobic microbes prevail in the sediment. Eutrophication is one of the major problems in the brackish Baltic Sea and its sub basin the GoF (Pitkänen 1994, Pitkänen et al. 2003). Seasonal oxygen depletion usually occurs after the decay of settled spring algal blooms. In the GoF, the formation of regional halocline and seasonal thermocline prevent the transport of oxygen to deep water layers leading to yearly anoxic periods in the bottom habitat (Haahti and Kangas 2006, Leppäranta and Myrberg 2009, Liblik and Lips 2011, Väli et al. 2013). During these periods, large amounts of P are released from the sediments to the water counteracting decreases in the external P load to the GoF (Pitkänen et al. 2001). Vertical mixing of the water body – due to autumnal turn over, stormy winds and salt–water pulses – eventually enhances oxygen conditions at the bottom but also brings P up to the water mass (Haahti 3- and Kangas 2006). In the water body the increased PO4 supply may accelerate the growth of harmful N–fixing cyanobacteria (Graneli et al. 1990, Elmgren and Larson 2001, Vahtera et al. 2010). Fe–bound and organic P are recognised to be the major mobile P pools in the surface sediments, but the Fe–bound P is considered to be in major control of the short term changes in the P–supply in the near bottom water (Lukkari et al. 2009b, Mort et al. 2010, Jilbert et al. 2011, Rydin et al. 2011). It has been estimated, that the concretions on bottoms in the eastern GoF alone may contain about 175 000 tons P (Vallius et al. 2011), which is ten times more than the 17 000 tons pool of P bound to Fe in the surface sediments of the accumulation bottoms (Lehtoranta 2003). Anoxia may occasionally reach the areas and water depths of abundant deposits of concretions (Pitkänen et al. 2003, Haahti and Kangas 2006, Hansson et al. 2011, Väli et al. 2013). Arsenic is present at a rather low level (mean 14 mg kg-1) in the sediments of the GoF (Vallius 2012). The World Health Organization (WHO) has set a current recommended limit of As in drinking–water 10 μg L-1. Its toxicity in nature is related to its oxidation state and solubility (Hughes 2002, Bissen and Frimmel 2003). Binding of As to Fe3+ oxides (Farquhar et al. 2002) reduces its bioavailability. Anoxia may affect the pool of Fe–bound As in the GoF sediments, and lead to its dissolution. Arsenic has been used as an indicator of leakage of toxic substances from the chemical munitions dumpsite in the Baltic Sea (Garnaga et al. 2006).

1.3.2. Bioremediation of harmful metals

Environments with high Fe and Mn concentrations may also be enriched with other, also harmful, metals and metalloids. Known Fe oxidisers, Gallionella and Sideroxydans, have evolved resistance against hazardous metals (Emerson et al. 2013). Many bacteria transforming Fe and Mn are also capable of catalysing the redox transformations of other transition elements such as radionuclides, and this way detoxify harmful elements. Dissimilatory bacterial reduction of uranium (U), chromium (Cr), cobalt (Co), gold (Au), palladium (Pd), technetium (Tc), vanadium (V) and selenium (Se) has significance for

19 precipitation of the elements and thus removing them from contaminated waste water and other sites (Lovley 1995, Lovley and Coates 2000, reviewed by Lovley et al. 2004). An addition of acetate to U–contaminated groundwater enhances the precipitation of U by bacterial reduction (reviewed by Lovley et al. 2004). Fe–reducing bacteria may 2+ outcompete SO4 –reducing species in sediments and thus prevent the production of toxic sulphides (reviewed by Lovley et al. 2004). Undesired effects of reduction include mobilisation of Fe and Mn, and co–precipitated P, As, and aluminium (Al) in aquifers which are sources of drinking water (reviewed by Dassonville and Renault 2002 and Lovley et al. 2004). Some bacteria may reduce As5+ to gain energy or to remove it from their cells (Anderson and Cook 2003, Akerman et al. 2011). This also leads to dissolution of As3+. Fe and Mn oxides cause harm by clogging plumbing systems. On the other hand, oxidised Fe and Mn species adsorb and coprecipitate metals and other compounds of environmental significance. Microbes produce poorly crystalline Fe oxides which are characterised by high surface area, reactivity and ability to be reduced (reviewed by Emerson et al. 2010). For example, the cadmium (Cd) adsorption ability of Mn oxides is high (Dong et al. 2003), and the oxidation–reduction cycle of Fe regulates the release of 3- PO4 (an important nutrient) from benthic sediment (e.g., Gunnars and Blomqvist 1997, Gunnars et al. 2002). Sequestration of P or As is enhanced by Fe3+ minerals produced also - by Fe–oxidizing bacteria (Hohmann 2010, Rentz et al. 2009). Microbial NO3 reduction coupled to Fe2+ oxidation can enhance the adsorption of As on the formed Fe–oxides as a bioremediation strategy even in anoxic system (Sun et al. 2009). In the presence of As5+ 2+ and SO4 , Desulfotomaculum auripigmentum can form insoluble arsenic trisulphide by 2+ respirative reduction (reviewed by Raab and Feldman 2003). SO4 reduction produces - hydrogen sulphides (H2S, HS ) which chemically reduce Fe and Mn oxides and form insoluble Fe sulphides (FeSx). This may inhibit diffusion of otherwise toxic sulphides from sediment to overlying water. The pollution of the Baltic Sea is a serious problem. Even if the situation concerning heavy metals in the sediments of the GoF is improving in general, still there is a drawback on Cd (Vallius 2012). Concretions can be considered as natural traps scavenging some toxic elements from the surrounding water (Zhamoida et al. 2007). The use of Fe–Mn concretions to monitor the input of heavy metals (Zn, Pb, Cd, Cu) into the Baltic Sea has been suggested, but it is not known what proportions of heavy metals entering the Baltic Sea are scavenged by the concretions (Glasby et al. 1997).

1.3.3. Oil–degradation potential in the sediments

The Baltic Sea is under environmental pressure of dense traffic and over 160 million tonnes of oil products is transported via the GoF, where the risk of a major oil accident exists due to increasing volume of oil transport (Helcom 2012). Constant exposure to small oil spills from ships may enrich petroleum hydrocarbon–degrading groups within the

20 bacterial communities. Microbial diversity and the potential of microbes to degrade oil compounds are widely studied in soils and in the Oceans (reviewed by Leahy and Colwell 1990). However, the ability of the microbial community to degrade petroleum hydrocarbons is poorly known in cold and brackish marine systems such as the Baltic Sea. In the water body of the ocean, Gammaproteobacteria are most often responsible for the degradation process of petroleum hydrocarbons for instance in the Gulf of Mexico (Hazen et al. 2010, Gutierrez et al. 2013), the North Sea (Brakstad and Lodeng 2005) and in the Mediterranean (Cappello et al. 2007). On the other hand, in the Baltic Sea, the dominant bacterial phyla responsible for the degradation of oil seem to differ along the salinity gradient: in lower salinity Betaproteobacteria and Actinobacteria are enriched (Reunamo et al. 2013), while in the higher salinity Alpha– and Gammaproteobacteria take over (Viggor et al. 2013). Anoxic areas are increasing worldwide due to eutrophication and changes in stratification properties of water in the coastal areas (Diaz and Rosenberg 2008). The coastal waters of the GoF meet regularly anoxia and bacteria may switch to alternative electron acceptors, or give way to a different bacterial community to take over in petroleum hydrocarbon degradation. Most of the known bacteria with the capacity to degrade saturated and aromatic hydrocarbons in the absence of oxygen are Proteobacteria (Widdel and Rabus 2001). Sinking of the oil tanker Prestige near the northern Spanish 2+ coast resulted in the increase of SO4 –reducing Desulfobacteraceae (Deltaproteobacteria class) in the anoxic sediment layers (Acosta–Gonzalez et al. 2013). Anaerobic degradation rates of petroleum hydrocarbons are much slower than aerobic rates, but the mineralised amounts of the substrate can be significant, from over 50% to undetectable level (reviewed by Leahy and Colwell 1990). Complex structures of polycyclic aromatic hydrocarbons (PAHs) are comparatively persistent especially in low–oxygen conditions. These compounds are harmful to biota, but microbes are able to degrade them (reviewed by 2+ - Bamforth and Singleton 2005). Microbial SO4 – (Aeckersberg et al. 1991) and NO3 – (Leduc et al. 1992, Alain et al. 2012) reduction processes are responsible for the bulk oxidation of aromatic hydrocarbons in anoxic sediments (Coates et al. 1997, Dell'Anno et al. 2009). It has been measured that in air–saturated water (8.6 mg/ L O2 at 25 °C) no more 2+ - than 2.8 mg/ L toluene can be degraded, whereas SO4 and NO3 are more soluble allowing possibly the anaerobic oxidation of greater amount of toluene (reviewed by 2+ Widdel and Rabus 2001). In SO4 –reducing conditions, through series of carboxylation and hydrogenation, also naphthalene is eventually metabolized to carbon dioxide (reviewed by Bamforth and Singleton 2005). Mn reduction has also been involved in petroleum hydrocarbon degradation in cold climate groundwater in the northwest Canada (Yeung et al. 2013). The bioavailability of Fe may limit its use as electron acceptor in the Ocean (Acota–Conzalez et al. 2013). Fe reducers, e.g. Geobacter, may degrade aromatic hydrocarbons such as benzene, toluene and phenol, which are difficult for other bacteria to be utilized as energy source in anoxic conditions (Lovley et al. 1989, reviewed by Lovley et al. 2004, Zhang et al. 2012). The rate of contaminant degradation coupled to Fe3+

21 reduction can be stimulated by adding chelators, which make Fe3+ more accessible, and electron–shuttling humic substances for the bacteria (reviewed by Lovley et al. 2004).

1.4. Aims of the study

The aims of this study were x to investigate bacterial community structure in Fe–Mn concretions of the Gulf of Finland Hypothesis: Bacterial community consists mainly of metal–oxidising and –reducing species. x to test if bacteria play a large role in the precipitation and dissolution of the Fe–Mn concretions and whether these processes can be controlled by the ambient environmental conditions Hypothesis: Concretion bacteria are able to oxidise and reduce Fe and Mn, and this way they contribute to the precipitation and release of elements of the concretions. Concretions in other environments may have similar contribution to the cycling of elements. x to test whether the concretion bacteria play a large role in the degradation of petroleum hydrocarbons and / or inactivation of harmful elements and whether these processes can be controlled by ambient environmental conditions. Natural recovery potential of the concretion bottoms and sediment in the GoF was tested. The implications of the microbial processes in the bioremediation were assessed. Hypothesis: Bacteria in the concretions are able to mineralise petroleum hydrocarbons in oxic and / or anoxic conditions. Bacteria in concretions responsible for degradation of petroleum hydrocarbons are metal reducers.

2. Materials and methods

2.1 Study sites and sampling procedure

The concretions and sediment used in this study were collected with a van Veen grab sampler in the Gulf of Finland, the Baltic Sea, during cruises by R/V Muikku and R/V Aranda (Table 1). The samples were stored in air–penetrating plastic bags, or in aerated aquariums in near–bottom sea water, or artificial brackish sea water constant at 5 °C, in dark room.

Table 1. The sampling sites of Fe–Mn concretions and sediment from the Gulf of Finland Sampling site and year Water depth Characteristics of the Paper (m) concretions and sediment 2007 Gulf of Finland 59 Spherical 0.5–2 cm in I N60° 01.00, E26° 04.85 diameter

2010 Finnish Estonian EEZ Spherical 0.5–2 cm in I, II area diameter station 1: 62 N59° 57.43, E25° 12.76 station 12: 54 N59° 57.24, E24° 59.58 station 17: 51 Spherical 0.5–2 cm in II N59° 57.59, E24° 51.05 diameter

N60° 9.22, E26° 36.42 55 Mixture of spherical and III N60° 10.55, E26° 45.29 48 crusts N60° 10.63, E23° 52.44 49 N60° 11.48, E26° 56.18 44 N61° 11.2, E18° 13.81 72 station LL4A: 59 Spherical and crust N60° 01.01, E26° 04.81 concretions Sediment was clay

2.2. Experimental set ups and methods

To investigate the role of bacteria in the formation of concretions, liquid and gradient cultures were used to test which bacterial groups were enriched in oxidative conditions accompanied by reduced Mn and Fe (Paper I). Negative controls were left without the inoculum. To investigate the release of Fe, Mn, P and As from the concretions, they were incubated in deoxygenated microcosms at 5 °C, 10 °C and 20 °C, in the dark, simulating anoxic conditions often prevailing in the GoF (Paper II). The media did not contain added Fe, Mn, P and As so these elements found in dissolved form originated from the concretions. Carbon source was left out to test autotrophy of the bacteria in all the experiments. Autoclaving was used as sterilization method in all the experiments. Summary of the methods is presented in the Table 2. To investigate the recovery potential of the concretion bottoms and sediments in case of a crude–oil spill in the GoF, three experimental set ups were constructed (Paper III). In the first experiment, mixtures of concretions were incubated with 10 mg Russian crude oil in oxic and anoxic microcosms. A negative control contained only brackish water

23 and crude oil without concretions. In the second (oxic) and the third (anoxic) experiments, LL4A concretions or sediment were incubated with 3 mg crude oil and 4.5 μL naphthalene. There were also autoclaved controls for all the experiments. Table 3 summarizes the set– up of the experiments covering crude oil with PAH degradation and naphthalene mineralisation.

Table 2. Summary of the used methods and analyses marked with the numbers of the papers where the details are described

Method Analysis Described and used in Paper Physical and chemical

x Freeze drying x Dry matter percent I, II, III x Ingition at 550 °C x LOI II x ICP–OES and ICP– x Fe, Mn, P and As I, II, III MS content of the concretions and liquid experiments(1 x Ferrozine method x Dissolved Fe2+ II, III (2 x Absorbance with x NH4 –nitrogen II indophenol at 630 nm x Infrared detection of x DIC and DOC(3 II CO2 x GC–FID and x Crude oil and total III chromatograph with PAHs analysis(4 MS x Radiorespirometry x Mineralization of III 14C–naphthalene(5 Microcosms

x Liquid batch culture x Enrichment of Mn– I with Mn2+ oxidising bacteria x Oxygen gradient x Enrichment of Fe I tubes with Fe2+ or and Mn –oxidising Mn2+ bacteria x Anoxic incubation x Microbial reduction II of concretions with of Fe, Mn, P and As and without added C in the concretions x Aerated and anoxic x Disappearance of III incubations of crude oil with concretions with 10 concretions mg crude oil x Aerated and anoxic x Disappearance of III

incubations of crude oil and PAHS concretions and with concretions or sediment with 3 mg sediment crude oil x Aerated and anoxic x Mineralisation of III incubations of naphthalene by concretions and concretion or sediment with 14C sediment bacteria naphthalene Microbial community analysis

x Phase contrast x Bacterial cell I microscopy morphology

x Epifluorescence x Enumeration of I microscopy (DAPI bacteria dye) x FastDNA Spin Kit x DNA extraction I, III for Soil x Wizard SV Gel and x DNA purification I, III PCR Clean –Up System x DGGE x Bacterial I, III community differences x TOPO TA Cloning x Cloning partial I, III Kit for Sequencing bacterial r16S gene x qPCR x Quantification of III PAH–RHDα GN and GP genes x Enzymes Hae III x RFLP I and Msp I x Sequencing service x Sequencing partial I, III bacterial r16S gene(6 Computational analyses

x Unweighted, x Similarity between I average clustering DGGE sequences (UPGMA) by profiles Quantity One Quantitation software x Ribosomal Database x Phylogenetic tree I Project (RDP) Treebuilder x DECIPHER x Chimera III

identification for 16S rRNA sequences x RDP taxonomic x Classification of I, III hierarcy classifier cloned sequences x Mothur x Determination of III OTUs(7 x The R software x Venn diagram and III version 3.1.0 rarefaction curves(8 x Excell, Microsoft x Basic statistical II, III analysis: means, standard deviations, correlations and their significances x SPSS version 22, x Differences between III IBM Corp. Released variances (ANOVA) 2013

Analyses: 1Timo Sara–aho SYKE, 2,3laboratory of SYKE, 4Anne Markkanen SYKE, 5,7,8Anna Reunamo, 6Institute of Biotechnology Helsinki (Finland) and Macrogen Seoul (South Korea)

Table 3. The set–ups of crude–oil with PAH–degradation and naphthalene–mineralisation experiments with concretions and sediment from the sampling station LL4A. AC= autoclaved control Flask Artificial Crude– 14C–labelled Sediment Concretions Nutrient AC brackish oil naphthalene addition water addition or (3 mg) Non–labelled naphthalene 1 x x x x 2 x x x x 3 x x x x 4 x x x x 5 x x x x 6 x x x x 7 x x x x x 8 x x x x x 9 x x x x x 10 x x x x x 11 x x x x x 12 x x x x x 13 x x x x x x x 26

14 x x x x x x x 15 x x x x x x x 16 x x x 17 x x x Flasks 16 and 17 were extra controls to test if naphthalene mineralization starts without oil addition.

3. Results

The results of this thesis are presented in three articles with tables and figures. Synthesis of the main findings is presented in this chapter.

3.1 Physical and geochemical characteristics of the GoF concretions

Spherical concretions had zoned inner structure of brown Fe and black Mn layers. The molar ratios between Fe and P, Fe and As as well as between P and As had significant positive correlations, whereas the correlation was significantly negative between Fe and Mn, Mn and P as well as between Mn and As (Tables 4, 5).

Table 4. Concentrations of the analyzed main elements and their molar ratios in the GoF concretions (Paper II) Element g kg-1 DW ± SD mol kg-1 ± SD Element Molar ratios relations

Fe 232 ± 42 4.2 ± 0.8 Fe:Mn 1.1–6.5 Mn 112 ± 36 2 ± 0.7 Fe:P 5.2–6.7 P 21 ± 3.5 0.7 ± 0.1 Fe:As 811–1177 As 0.312 ± 0.054 0.0042 ± 0.0007 P:Fe 0.12 As:Fe 0.0007

Table 5. Pearson correlation coefficients “r” of the analysed elements of the GoF concretions (Paper II) Element Fe p–value Mn p–value P p–value 2–dimensional

Fe Mn -0.92 <0.0005 P 0.835 0.005 -0.9 <0.001 As 0.737 <0.05 -0.855 <0.005 0.961 <0.0001

Concretions had high content of water and organic C, and they were colonised by millions of morphologically and taxonomically varying bacteria (Table 6, Papers I, II, III).

Table 6. Physical characteristics and microbial density of the GoF concretions (Papers I, II) Dry matter content (% ± SD) Loss on ignition Microbial cell number (LOI%) (cells g-1 DW) 50 ± 5 12 6.7 x 107

3.2 Fe and Mn –oxidising and –reducing processes in the GoF concretions

Microbes from the concretions enhanced the oxidation of Mn in the liquid cultures in a week and in the semisolid gradient tubes in a month (Figure 1, Paper I: Figures 3, 4). Mn oxidation accelerated after every transfer of inocula to fresh gradient–culture tubes, especially when they had the added C source indicating the enrichment of Mn–oxidising species in the experiments (Figure 1, Paper I: Table 4). Dark precipitation was not observed with the autoclaved inocula or without the inocula. In the liquid cultures uniform cells were clustered around granule–like Mn oxides (Paper I: Figure 5) whereas in the semisolid media bacteria appeared mainly club– formed or rod–shaped (Figure 1, Paper I: Figure 4b).

Figure 1. (upper panel) Black Mn–oxide precipitate formed around the inocula, originating from Fe–Mn concretions, in the opposing gradients of diffusing MnSO4 from the bottom plug and O2 from the air space. First three tubes from the left have C source, the last three ones from the right do not have C source. (lower panel) Club–formed bacteria growing attached to Mn oxides in the gradient culture tubes. (Paper I)

Microbial growth was observed as reddish–brown layer around the original inoculum in FeS–containing gradient tubes with or without the added C. Bacterial cells were commonly filamentous and attached to Fe–oxide particles (Figure 2, Paper I: Figure 7a, b). Similar layering was not observed with the autoclaved inocula, or without the inocula. The growth–layer formation accelerated after every transfer to fresh gradient tubes.

Figure 2. (upper panel) Formation of Fe–oxide disks in the opposing gradients of diffusing

FeS from the bottom plug and O2 from the air space. The tube in the middle is not inoculated. First three tubes from the left have C source, the last three tubes from the right do not have C source. (lower panel) Filamentous bacteria growing attached to Fe–oxide particles in the gradient culture tubes. (Paper I)

In the anoxic microcosms concretions released dissolved Fe, Mn, P and As at every incubation temperature if they were supplied with the added C source (Table 7, Paper II). The released concentrations of Mn always exceeded those of Fe (Table 7, Paper II: Figure 4a). Additionally, released Fe reached its highest values later than the dissolved Mn (Table 7, Paper II: Figure 4b). P and As were released from the concretions with Fe (Paper II: Figure 5a). Twice more P (P:Fe 0.2 mol/ mol) was released from the concretions than was expected on the basis of the P:Fe ratio in the solid concretions (Table 4). The As:Fe ratio was around the same in the intact concretions and at dissolved form at 5 °C (Table 4), but at higher temperatures the ratio increased to 0.0012 which was explained by the disappearance of Fe from the solution at 10 °C and 20 °C (Paper II: Figure 4b, 5b). Also + Mn disappeared from the solution at 20 °C. The consumption of DOC and NH4 were in + line with the production of DIC. There was NH4 consumption below the detection limit also in the incubations without the added C (Paper II: Figure 3). + Autoclaved controls showed only slight consumption of the added NH4 and DOC, some increase in DIC and minor dissolution of Fe, Mn, P and As. The values were, however, below the values measured from the non–sterile incubations.

Table 7. Concentrations of released Fe, Mn, P and As and their release rates from the concretions of the GoF during anoxic incubations (Paper II) Element The highest Release rate ranges Release rate range Temperature released in the incubations (μmol m-2 d-1) (°C) concentrations (μmol g-1 DW estimated for in the concretion d-1) at bottoms with incubations different average 15 kg (μmol g-1 DW temperatures concretions m-2 in concretion) at the GoF at 5 °C different temperatures, and time Fe (weeks) 5 4.9 14 0.01–0.12 84–1500 10 5.5 8 0.02–0.61 20 3.3 7 0.02–0.22 Mn 5 8.4 6 0.002–0.42 30–6300 10 7.5 6 0.003–0.24 20 6.3 5 0.14–0.31 P 5 0.9 15 0.01–0.02 190–550 10 0.4 5 0.003–0.05 20 0.95 5 0.003–0.05 As 5 0.004 15 0.00002–0.0002 0.2–3.3 10 0.002 5 0.00003–0.0002 20 0.006 5 0.00001–0.0003

3.3 Disappearance of crude oil and mineralisation of 14C naphthalene

In the range of 59–70% of the added crude oil disappeared in the oxic concretion microcosms (Paper III: Table 2). In the anoxic microcosms correspondingly 51–61% of the added crude oil disappeared. In comparison, in the sediment microcosms 54% and 41% of the crude oil disappeared in oxic and anoxic conditions, respectively. Nutrient addition did not enhance the degradation significantly (Paper III: Table 2). The crude–oil disappearance rate calculated from the mixed concretion incubations was faster in oxic than anoxic conditions during the first weeks, but the rates stabilized for the both treatments after six weeks of incubation (Paper III: Figure 1). After seven weeks of incubation the crude–oil concentration dropped suddenly, 46% also in the aerated flasks containing autoclaved mixed concretions, and 80% after eleven weeks in the aerated flasks containing autoclaved LL4A concretions with sediment, but only 10% in the anoxic autoclaved controls. It is noteworthy that there was no corresponding decrease in the aerated flasks containing only crude oil and artificial 31 brackish sea water. Based on these results the proportion of biological crude–oil degradation may have been at least 13% in oxic microcosms with mixed concretions. In the anoxic microcosms the proportion of biotic degradation would have been at least 5% with mixed and ranging 25–50% with LL4A concretions and around 30% with the sediment. Vast majority (96%) of the analysed PAH compounds diminished in all the microcosms containing LL4A concretions or sediment in oxic and anoxic conditions. The disappearance of 79% of PAHs in the aerated and 35% in the anoxic autoclaved controls was problematic. Based on these values, the proportion of biological PAH degradation would have been about 17% in oxic and 60% in anoxic incubations. The disappearance patterns of PAHs were different between the autoclaved and non–sterile microcosms. The 2– and 3–ringed PAHs had mostly disappeared from the microcosms whereas in the autoclaved controls these compounds remained (Paper III: Figure 2). Mineralization of 14C–naphthalene was significantly higher in the oxic LL4A concretion incubations than in the anoxic incubations being 5–8% and 2–4%, respectively (Paper III: Figure 3, Table 2). In comparison, in the oxic sediment incubations 5–14% of naphthalene was mineralized, but in the anoxic incubations only < 3%. Autoclaved controls showed only 0.2% mineralization of naphthalene. The additions of crude oil led to <200 μg L-1 dissolution of Fe and <3 mg L-1 Mn when LL4A concretions and sediment were incubated. But, Fe and Mn dissolved with great variation between triplicate units containing mixed concretions (Paper III). Concentrations of dissolved Fe (Fe2+ was 90% of total Fe) ranged between 93–800 μg L-1 in anoxic treatments (Figure 3). On the contrary, the concentrations of Mn were up to hundred times higher than Fe, varying between 1–74 mg L-1 (Figure 4), but there was no significant difference between the oxic and anoxic treatments.

Figure 3. Dissolved Fe from the mixed concretions incubated with crude oil in oxic and anoxic conditions. AC indicates autoclaved control.

Figure 4. Dissolved Mn from the mixed concretions incubated with crude oil in oxic and anoxic conditions. AC indicates autoclaved control. (Paper III)

Autoclaving also released Fe and Mn into the solution, and is therefore not asufficient method to form chemical control incubations.

3.4 Bacterial communities of the GoF concretions

Bacterial communities of the surface and the interior of the concretions differed according to clustering analysis of DGGE bands (Paper I: Figure 1). About 9% of the OTUs were shared between the initial LL4A concretions and sediment, but only 2.3% of the OTUs were shared between the initial mixed concretions and LL4A sediment (Paper III). This may be explained by the fact that the LL4A concretions and sediment had been collected from the same sampling site and stored together. Around 30–40% of the cloned sequences belonged to unknown taxa (Papers I, III). Half of the cloned sequences were affiliated with Proteobacteria and 10–20% were distributed among 12 phyla (Figure 5, Papers I, III).

Figure 5. Phylogenetic affiliation of the bacteria in the clone library of Fe–Mn concretions from the GoF. n= 114 sequences (Paper I)

All the common Proteobacteria classes were represented in the clone libraries, and their distribution differed between the concretions and sediments (Paper III: Figure 5). Gammaproteobacteria were the most dominant in the concretions, whereas Deltaproteobacteria were in the sediments. The proportion of unclassified Proteobacteria was highest in the concretions. Distribution of the bacterial phyla differed between the communities of the concretions and the surrounding sediment (Paper III: Figure 4).The cloned sequences were compared to their closest matches in the GenBank of sequences and they could be categorized into ecotypes based on the environments they were obtained from (Paper III: Figure 6). Some 20–30% of the bacterial sequences from the intact concretions were similar to the sequences earlier found from the Baltic Sea concretions (Paper III: Figure 6, Table 3) which, in turn, were closely related to the uncultured clones of Gamma– Delta– and Betaproteobacteria from sea floor lavas (Santelli et al. 2008), Baltic Sea sediments (Edlund et al. 2008), oil–contaminated sediment (Winderl et al. 2008) and metal–rich environments (e.g., Nemergut et al. 2004) (Paper I: Figure 2). The unclassified sequences had high similarity scores to uncultured bacteria originating, e.g., from sediments, hydrocarbon–polluted soil and uranium mine (unpublished). Sediment samples showed a distinct pattern, different to concretions. Only few clones from the sediment experiments matched to the sequences previously obtained from Fe–Mn concretions of the Baltic Sea, but nearly 40% of the clones matched to the sequences previously found from sediments (Paper III: Figure 6, Table 3).

3.5 Mn– and Fe–oxidising bacteria enriched from the concretions

Different bacterial species and strains were enriched depending on the cultivation media with reduced Fe or Mn, agitation and addition of C (Paper I: Table 4, Figure 6). The closest matches for the cloned sequences from the Mn–liquid media were to cultured strains of Pseudomonas, Bacillus and Sphingomonads. The closest matches for the cloned sequences from the Mn–gradient culture were to different Pseudomonas sp., Prosthecobacter sp. (Verrucomicrobia), Rheinheimera sp. and Comamonadaceae. The Fe– gradient culture favoured the enrichment of Shewanella baltica, Thalassolituus oleivorans, Pseudomonas sp. and Comamonadaceae.

3.6 Bacterial community responses to crude–oil exposure

After incubation of concretions and sediment with the added crude oil, the proportion of unclassified bacteria decreased and the proportion of Proteobacteria increased (Paper III: Figure 4). Alpha– and Betaproteobacteria were enriched in the concretions, Gamma– and Betaproteobacteria in the sediments (Paper III: Figure 5). The proportion of Nitrospira increased in the anoxic treatment of mixed concretions. The distribution of clones between the different environments of origin (ecotypes) remained more or less the same after the crude–oil exposure, however, the closest matches were different (Paper III: Figure 6, Table 3). After oxic incubation of concretions, the proportion of the closest matches originating from oil–contaminated sites, and metal effected environments such as Mn–rich sediments and deep–sea polymetallic nodules increased slightly (Paper III: Figure 6). After anoxic incubations of mixed concretions with crude oil, clones classified as Albidoferax (Betaproteobacteria, Comamonadaceae) (Ramana and Sasikala 2009) were enriched and their closest matches were found to uncultured cloned sequences from a pyrite mine (Hao et al. 2012). The clones classified as Nitrospira (Paper III: Figure 4) matched to sequences originating from various environments: Antarctic, As–contaminated sediment, Fe and S–precipitating microbial mat (Omoregie et al. 2008) and lava. After oxic incubation of sediment with the added crude oil the proportion of clones matching to sequences originating from sediment and petroleum hydrocarbon contaminated sites, as well as from water or waste water origin, increased. Clones classified as Gammaproteobacteria, similar to sequences originating from oil contaminated sediment (Paisse et al. 2010) and Baltic Sea sediment enriched (Edlund et al. 2008). After anoxic incubation of sediment with the crude oil there was also increase of clones which matched to sequences originating from metal effected environments such as Mn–rich sediment (Vandieken et al. 2012). Three clones were affiliated with cultured species Geopsychrobacter electrodiphilus (Deltaproteobacteria) (Holmes et al. 2004). Mixed concretions were incubated with crude oil, but without the added naphthalene. The copy number of gram positive PAH ring–hydroxylating dioxygenase (GP

PAH–RHDα) seemed to increase during the first four weeks of the experiment in both the oxic (1.8×105 ±9.4×104) and the anoxic (1.3×105 ±1.2×105) incubations, but because of the large variation no significant difference was observed. At the end of the experiment the gram negative (GN) and GP PAH–RHDα gene copy levels were around the same level as in the initial samples (Paper II: Table 2). In the oxic LL4A concretion microcosms, the number of GN PAH–RHDα gene copies increased significantly 1000–fold during the experiment (Paper II: Table 2). The number of GN PAH–RHDα gene copies was significantly higher in the concretion incubations than in the sediment incubations. In the anoxic experiment, the number of GN PAH–RHDα gene copies also increased significantly in the concretion microcosms, but no rise was detected with the GP gene copy numbers. In comparison, in the oxic LL4A sediment incubations the number of GN PAH–RHDα gene copies increased significantly four orders of magnitude in the microcosms with nutrients addition.

4 Discussion

4.1. Concretions as microbial habitats in the GoF

The spherical concretions collected from the GoF were fragile and their high water content indicated that they were porous inside (Papers I, II). Gradients of dissolved oxygen (O2), metals, nutrients and other substances may form varying micro–niches for bacteria on the pitted surface and in the porous interior of the concretions as well as in the interface between the concretions and surrounding sediment. Deep–sea Mn nodules also have a large connected pore space (Blöthe et al. 2015). Fe oxidises quickly by O2 but diffusive barrier may form between concretions and sediment where microaerophilic bacteria may oxidise Fe (Paper I). Even if direct Mn oxidation is slow, it is greatly enhanced by bacteria (Paper I, see also Tebo and Emerson 1986). Biofilms may develop in and outside Fe and Mn oxides in the sediment, which may be colonised by metal–reducing and –oxidizing bacteria (Stein et al. 2001). Cluster analysis of the DGGE bands showed that at least the bacterial communities of the surface and the interior of the concretions differed from one another (Paper I). Difference between the communities of the interior and exterior has also been discovered in nodules sampled from the central South Pacific (Tully and Heidelberg 2013) and from the north–east Pacific (Blöthe et al. 2015). The number of microbial cells in the GoF concretions was within the range counted from the biofilms in Fe–Mn mineral crusts in Tennessee River basin caves (Carmichael et al. 2013) and from the deep–sea Mn nodules (Blöthe et al. 2015) but ten times higher than in soil Fe–Mn nodules in China (Zhang et al. 2014). The high number of bacteria in the GoF concretions was in contrast with the assumption of concretions being selective and unfavourable habitats (Papers I, III). Organic C content of the concretions was within the range measured from the surface sediments of the GoF (7–23%, Lehtoranta 2003, Lehtoranta et al. 2004). It is 36 possible that Fe–oxides bind organic matter (McKnight et al. 1992, Eusterhues et al. 2010, Sharma et al. 2010, Riedel et al. 2013) in the concretions during their growth, but it has not been known to what extent heterotrophic bacteria are able to utilise this C (Paper II). Organic C has been inferred to be one of the most significant variables for structuring the metabolically active bacterial communities along the vertical gradient in the Baltic Sea sediments (Edlund et al. 2008). In the GoF the concentrations of Fe, Mn, P and As are several times higher in the concretions than in the surrounding sediments or settling particulate matter (Paper II: Figures 7a,b, 8a,b, 9). Concretions and settling matter seem to cumulate Mn in the GoF (Leivuori and Vallius 1998, Paper II: Figure 7a, b). Therefore, concretions are distinct metal–rich habitats, which may create selective pressure on the formation of the microbial community, different from the ones in the surrounding sediment (Paper III).

4.2. Bacterial community structure of the Fe–Mn concretions in the GoF

The cloned sequences were compared to their closest matches in the GenBank and categorized into ecotypes based on the environment they were obtained from (Paper III: Figure 6). The pattern of matching original environments was strikingly similar for all concretion samples despite their original sampling site, but different from the surrounding sediment. The primers used for cloning and sequencing had a broad coverage revealing many phyla of the concretion bacteria, but a high proportion of unclassified bacteria seems to be characteristic to concretion bacterial communities (Paper I: Table 3, Figure 2, Paper III: Figure 4) as also to sediment communities in the Baltic Sea (Edlund et al. 2008). Even if phylogeny cannot always explain metabolism, the concretion bacteria were distributed among thirteen phyla, mostly Proteobacteria, indicating metabolic variability. A large proportion of the closest matches were to sequences obtained from the concretions themselves (Paper III: Figure 6, Table 3). The high frequency of these clones were affiliated with specific Alpha– Beta– Gamma– and Deltaproteobacteria which may indicate that there are new lineages of bacteria within the Rhizobiales and Rhodospirillales of the Alphaproteobacteria, Burkholderiales of the Betaproteobacteria, Cromatiales of the Gammaproteobacteria, and the Syntrophobacteriales of the Deltaproteobacteria in the Baltic Sea concretions (Paper III: Table 3). This in turn indicates that concretions are distinct habitats, not only characterised by their physical and chemical properties but also by the bacterial community they host. Sediments have been inferred to be phylogenetically the most diverse environment type, and the major environmental determinant of microbial community composition has been salinity (Lozupone and Knight 2007). Salinity shapes the bacterial populations of Fe– mat communities, and some coexistence between marine and freshwater populations occur in brackish Sheepscot River estuary in Maine, USA (McBeth et al. 2013). Sediments around polymetallic nodules in the Pacific Ocean had more diverse bacterial community but only around 13% of the sequences were shared between the sediments and nodules

(Tully and Heidelberg 2013, Wu et al. 2013, Blöthe et al. 2015). The most abundant bacterial species detected only in the oceanic nodules were Mn–oxidising Pseudoalteromonas and Alteromonas (Gammaproteobacteria), and the nodules contained higher archaeal diversity than the sediment across all the sampling sites (Wu et al. 2013). A more recent study of deep–sea Mn nodules from the north–east Pacific revealed that their bacterial community was dominated by Shewanella sp. and Colwellia sp. (Gammaproteobacteria) not detected from the surrounding sediment (Blöthe et al. 2015). Separation of bacterial community structures was observed also between Fe–Mn concretions and their surrounding soil, and explained by different organic matter contents (He et al. 2008, Zhang et al. 2014). Nodule samples were characterised by high proportion of Betaproteobacteria whereas soil communities were dominated by Acidobacteria (Zhang et al. 2014). These examples from the Oceans and soils indicate that formation processes of metal–rich concretions in the specific environments, also the Baltic Sea, may have caused species sorting and resulted in the differences between the bacterial communities.

4.3. Bacterial Fe and Mn oxidation contribute to the formation of the GoF concretions

Concretion bottoms have formed in oxic conditions in the GoF. Especially abundant concretion deposits are located in the eastern GoF (Glasby et al. 1997) where the conditions most likely remain oxic year round due to turnover of the water body (Liblik and Lips 2011). Bacterial oxidation of Fe and Mn may partly explain the fast formation rate of the concretions in the Baltic Sea (Paper I). Concretions of the GoF were not dominated by known metal–cycling species, in contrast to a recent study which showed deep–sea Mn nodules being dominated by Mn–cycling Colwellia sp. and Shewanella sp. (Blöthe et al. 2015). The ability to oxidise or reduce metals remains unknown for the large proportion of the Baltic concretion bacteria. A relatively small proportion of the bacterial community might be active (Edlund et al. 2008) and responsible for metal cycling. Different bacteria were enriched in the liquid and semisolid Mn–oxidation experimental conditions (Paper I). Liquid media encouraged the growth of known Mn– oxidisers such as Pseudomonas (Ofek et al. 2009), Bacillus and Sphingomonadaceae (Graff and Conrad 2005) known to biodegrade contaminants (Zheng et al. 2011). Appendage–forming Prosthecobacter (Verrucomicrobia; Hedlund et al. 1997) and Rheinheimera baltica (Gammaproteobacteria) were enriched in the semisolid Mn– containing gradient, and they are also among the most abundant taxa of the bacterial community in the water column of the Baltic Sea (Brettar et al. 2002, Herlemann et al. 2011). There are no previous reports on their ability to oxidize Mn. Similar bacteria to hyphal Prosthecobacter have been observed in the enrichment cultures obtained from the surface layers of the concretions from the Western Baltic Sea (Ghiorse and Hirsch 1982). Semisolid media may have better simulated the natural conditions for slowly–growing bacteria on the surface of concretions.

Bacteria may grow and form diffusive layers where they are able to oxidise Fe. Despite their growth in the disk of Fe oxides (Paper I, Table 4), the bacteria strains did not fall into categories of known Fe oxidisers (reviewed by Weber et al. 2006). The candidatus class belonging to the neutrophilic Fe–oxidising bacteria Zetaroteobacteria (Emerson and Moyer 2002, Emerson et al. 2007) were not found from the concretions or the sediment of the GoF. On the contrary, Shewanella baltica (Ziemke et al. 1998) is a known heterotrophic Fe reducer capable to accumulate Fe and exhibit enrichment in microaerophilic conditions. Shewanella may have contributed to the reduction and release of Fe and Mn from the concretions in the anoxic microcosms (Paper II). Thalassolituus oleivorans is known to utilize petroleum hydrocarbons (Yakimov et al. 2004). Members of the family Comamonadaceae are found from various environments (soil, Morales et al. 2009) and enrichments, such as denitrifying waste water treatment systems (Lu et al 2014) and Fe–oxide producing microbial community (Anderson et al. 2006). Rather than being autotrophic, concretion bacteria may have been oligotrophic. Concretion formation processes are very beneficial considering the state of the GoF, because Fe and Mn oxides scavenge P, As and organic matter to the concretions (Paper II). An increase in the concentration of Fe seems to predict the increase in the concentrations of P and As in the concretions of the GoF and GoB (Paper II: Figures 8a, 9). Concretions in the Baltic Sea, Oceans, lakes and soils share some geochemical characteristics which may favour microbial oxidation and reduction depending on the environmental conditions, and affect element cycling locally (Papers I, II). Spherical concretions of the GoF have alternating brown Fe and black Mn layers, similar to those observed in soil concretions in Greece (Gasparatos et al. 2005) and in Fe–Mn crusts in south–west Baltic Sea (Marcus et al. 2004). Concretions in the GoF accumulate more Fe than those found from the Oceans, and more Mn than concretions found from soils (Ingri 1985, Paper II: Figure 7a). The comparable high Fe–bound P and As content of the concretions in the GoF suggests them being a potential and remarkable repository of these elements (Paper II).

4.4. Microbial Mn and Fe –reduction control P and As release from the concretions

Anoxia as such did not cause the release of metals from the concretions in the incubation experiments. The accumulation of dissolved forms of Fe and Mn and formation of black 2- precipitate of FeSx verified that the metals from the concretions and SO4 from the sea water were used as electron acceptors by the concretion bacteria, but only if they were supplied with labile C in anoxic conditions (Paper II). The mineralization of the added C seems to explain the release of P and As from the concretions as well (Paper II: Figure 6a, b). Also components of crude oil may be used as C source by heterotrophic concretion bacteria during anoxic conditions (Paper III). Mn oxides are reactive and known to degrade humic substances to lower molecular weight compounds which may be used for microbial growth (Sunda and Kieber 1994). But, it seems that the extent at which the organic material in the concretions was utilised by Fe and Mn –reducers is low, leaving

39 concretions a role of a stock of organic C (Paper II: Figure 2). Concretions may even + participate in the cycling of N (Paper II: Figure 3). Unexpectedly; NH4 was depleted to below the detection limit in the incubations where the only organic C available was bound to the concretions. However, minor amounts of metals dissolved, suggesting that some other anaerobic reactions took place. Even if the Fe concentrations are generally higher than Mn in the concretions, more Mn was dissolved from the concretions. The alternating oxide layers of Fe and Mn in the concretions may have been available for reducing bacteria in succession. Mn reduction may have yielded more energy for bacteria following the free energy gain, or bacteria may have even been specialized in Mn reduction (e.g., in Black Sea, Thamdrup et al. 2000). Known Fe–reducing species, such as Geobacter, belong to Deltaproteobacteria (Weber et al. 2006). However, initial concretions were not dominated by the known Fe and Mn – reducing bacteria (Paper I, III). Nevertheless, Shewanella baltica (Gammaproteobacteria) was enriched in microaerophilic O2–gradient tubes inoculated with concretion suspension (Paper I), Albidoferax (Betaproteobacteria) in anoxic microcosm with concretions and crude oil, and Geopsychrobacter electrodiphilus (Deltaproteobacteria) (Holmes et al. 2004) was detected in microcosms with surrounding sediment and crude oil (Paper III). These species are known for their ability to reduce Fe, and may have contributed to the dissolution of the metals. Fe2+ may also chemically reduce Mn3+/4+ and this reaction may have increased the concentration of dissolved Mn and hence inhibits the dissolution of Fe into the solution (reviewed by Lovley 1991, Postma and Appelo 2000). Mn oxides can even be reduced by As3+ (Oscarsson et al. 1982, Chen et al. 2006). These reactions can be difficult to distinguish from direct enzymatic reduction by microbes.

4.4.1. Effects of temperature on Mn and Fe –reduction and P and As release from the concretions

Fe and Mn reduction were dominant at 5 °C (Lovley and Phillips 1987) which is near the bottom temperature where concretions occur. At higher temperatures Fe disappeared from dissolved form with P, whereas Mn and As seemed to stay in the solution, except at 20 °C where also Mn possibly adsorbed onto the solids (Hamilton–Taylor et al. 1996). 2- At higher temperatures competing SO4 reduction may have taken over (Jensen et al. 1995, Knoblauch and Jørgensen 1999), leading to disappearance of Fe from the solution as black Fe–sulphide precipitates. Usually Fe–sulphides have got poor ability to capture P (Roden and Edmonds 1997), as in the sediments with low oxygen conditions. Nevertheless, very low concentrations of P were measured from the solution when the black precipitates were formed. The explanation may be that, in the concretions the concentration of Fe is high, and so Fe oxides may retain released Fe2+ and P in black– coloured suspension (Golterman 1995). It is also possible, but only speculative in our 2+ 3- experiments, that released P precipitates as Fe –PO4 (e.g. vivianite) when Fe reduction

2- exceeds SO4 reduction (H2S production). These conditions may hold e.g. in the Bothnian Sea (Slomp et al. 2013). There are also studies which propose that P may be associated with Mn (Mort et al. 2010, Jilbert and Slomp 2013). Usually FeS sequesters As, which can mitigate As mobilisation (Burton et al. 2014). Possibly As was bound to Fe–complexes which easily released As (Sharma et al. 2010) from the concretions. Some bacteria may use As5+ as an electron acceptor and reduce it to more soluble As3+ to gain energy or to remove this toxic element from their cells (Akerman et al. 2011, Anderson and Cook 2003). Dissimilatory As reduction may contribute significantly to As mobilization where the reducing bacteria are native to the sediments (e.g., Aberjona watershed U.S.A., Ahmann et al. 1997). Concretions incubated with crude oil in anoxic microcosms had bacterial clones similar to those observed in metal–rich As– containing environments but their ability to cycle As remains unknown (Paper II: Figure 6). Concretions host arsenic tolerating species such as Pseudomonas and Bacillus (Anderson and Cook 2004).

4.4.2. Release of Fe, Mn, P and As vs. sediment effluxes

The measured release rates of Fe, Mn and P may be higher from the concretion bottoms than the average fluxes from the GoF sediments during anoxic conditions (Paper II: Table 3). These experimentally derived values are within the release rate range earlier reported from the sediments (Jensen et al. 1995, Lehtoranta 2003, Lehtoranta and Heiskanen 2003, Lukkari et al. 2009a, b, Pakhomova et al. 2007). Mn and Fe fluxes out of sediment have been measured from short term (hours) anoxic chamber incubations in one study site which contained Fe–Mn concretions in the GoF (Pakhomova et al. 2007). Nearly one third of the Fe and Mn pool of sediments may be released during a month (Kristiansen et al. 2002, Kristensen et al. 2003). In our experiment the incubation time of the concretion microcosms was comparatively long (moths), but only 0.1, 0.4, 0.1 and 0.009 % of the total Fe, Mn, P and As in the concretions were released to the solution respectively at 5 °C. Concretions may then release only minor part of their Fe–bound P and As reserves slowly during prolonged anoxic period. Arsenic has not been considered the major contaminant of the GoF sediments (Vallius 2012) in contrast to, for instance, contaminated Balmer Lake sediment in Canada, where increased phytoplankton production enhanced anoxia and thus the release of As from sediment deposits (Martin and Pedersen 2002). As levels in the surface sediments of the GoF commonly stay around 0.37 μmol As g-1 at highest (Paper II: Figure 9, Zhamoida et al. 2007, Vallius 2012). Average As accumulation in the GoF sediments has been estimated to be 0.6 μmol m-2 d-1 (calculated from Leivuori and Vallius 1998), but the release rates from the concretions in anoxic conditions may temporarily be over five times higher (Paper II: Table 3). The released proportion of As from the concretions was small compared to anoxic sediment, where even 28% of As present can be mobilized, like in Aberjona watershed in the U.S.A. (Ahman et al. 1997).

During anoxic conditions, Mn oxides are reactive and can oxidise Fe2+ and As3+ (Zhang et al. 2014), which may precipitate back on the surface of a concretion. Newly formed Fe phases and remaining Fe–oxides may still retain As, even if metal–reducing bacteria dissolve Fe in anoxic conditions (Neidhardt et al. 2014). The same may eventually apply with concretions.

4.5. Degradation potential of crude oil and PAHs by Fe–Mn concretions in the GoF

Concretions from the different sampling sites and sediment (from LL4A) were incubated several weeks with crude oil and 14C–naphthalene to simulate the natural recovery of the GoF bottoms after a fresh crude–oil contamination under oxic and anoxic conditions (Paper III). The slowdown of the degradation rates during the mixed concretion incubations may have resulted from the quick degradation of light fraction leading to the slow degradation of complex residual compounds of the added crude oil. Usually in the case of an oil–accident, the light fractions of crude oil disappear since they evaporate and are degraded already in a water phase. This leads to sinking of more complex compounds – such as PAHs – to the sea bottom. During anoxic conditions these compounds may remain under sediment layers or be degraded very slowly. Therefore the results derived from our microcosms may be applied in case of a fresh crude–oil spill, or if the contaminated sediment with concretions was dredged and treated on land. Similar degradation potentials for crude oil were observed by concretions and sediment in both oxic and anoxic conditions, even if concretions and sediment are physically and chemically different environments for microbes; concretions being solid, but porous metal precipitates and sediment comprised mainly of fine grained clay containing metal oxides. Formerly, it has been shown to be more essential to have bacteria already adapted to oil contamination in sediments (Coates et al. 1997, Anderson and Lovley 1999) than to add Mn4+ (Li et al. 2011) or Fe3+ (Li et al. 2010) as electron acceptors for the bacteria at a polluted site (Coates et al. 1996a). The bacterial community in concretions must already have been adapted to higher metal concentrations than in the surrounding sediment (Paper II) and the bacteria can slowly adapt to oil degradation by addition of crude oil to their environment. Previously, addition of nutrients to enhance petroleum hydrocarbon degradation has been reported in many studies (Lee and Levy 1989, Pelletier et al. 2004, Coulon et al. 2007). The degradation is not likely to be limited by nutrients in the GoF, because the + concentrations of NH4 and phosphate in pore–water commonly exceed milligram levels (Lehtoranta and Pitkänen 2003). Sediment and concretions contain high amounts of P which may dissolve (Paper II), and become applicable for oil–degrading bacteria.

4.5.1. Fe and Mn reduction under the crude oil exposure

Different micro niches may have developed inside and on the surface of the concretions (Paper I) as well as in sediment layers, where bacteria may have used fractions of crude oil as their C source and Fe and Mn as electron acceptors during anoxic conditions (Papers II, III). Mn has been used as electron acceptor by bacteria in cold–climate groundwater system contaminated by petroleum hydrocarbons (Yeung et al. 2013), and in soil percolation columns (Langenhoff et al. 1996). Fe2+ could have dissolved due to bacterial reduction, and coupled to anaerobic mineralization of aromatic contaminants (Lovley et al. 1989, Lovley et al. 1994), as has been shown with naphthalene in sediments under Fe3+– reducing conditions (Anderson and Lovley 1999). During anoxic conditions also As may be released and cause harmful effects to the environment (Paper II). Microcosms with LL4A concretions and sediment behaved differently. Fe and Mn did not dissolve to same extent as from the mixed concretions in anoxic conditions. Concretions in these two experiments were from different sampling sites and differed from one another in size and shape. Nitrate was added as nutrient only in LL4A concretions and sediment and it may have acted as electron acceptor rather than Fe and Mn in these incubations (LaRiviere et al. 2001, Wang et al. 2012).

4.5.2 Abiotic and biotic disappearance of crude oil and PAHs

The high metal concentrations in the concretions may explain why the disappearance of crude oil and PAHs from the aerated autoclaved control incubations with LL4A concretions and sediment was as high as from the non–sterilised incubations (Paper III: Table 2). Fe–bearing minerals are reactive and can oxidise alkanes and other hydrocarbons under certain conditions (Surdam et al. 1993, Seewald 2003, Ferrarese et al .2008). Autoclaving may have altered the chemical structure of the concretions and sediment, and thus have led to chemical degradation of crude–oil fractions in the incubations. Autoclaving may have changed the adsorption properties of porous concretions and fine– grained sediment, because it can cause dissolution of organic C, P and Mn from soils and sediment (Tuominen et al. 1994). Also, the extraction method may have not been appropriate for this substrate. These explanations are supported by the stable concentrations of crude oil in the aerated control flasks containing only artificial brackish sea water and crude oil. In soil system, volatilization has accounted for 30% loss of naphthalene, and abiotic reactions 2–20% of the reduction in two– and three–ringed PAHs (Park et al. 1990). The indication of recovery of bacteria in the autoclaved samples was not observed when comparing the crude oil disappearance rates of autoclaved and non– autoclaved incubations of mixed concretions, even if the added crude oil may contain bacteria (Paper III: Figure 1). The disappearance patterns of PAHs differed between the autoclaved and the non– autoclaved microcosms supporting microbial degradation as the explanation (Paper III:

Figure 2). The ratio of 2–6 ringed PAH compounds had changed notably in the non–sterile incubations, but not in the autoclaved ones, indicating that in the non–sterilised incubations 2–, 3– and 4–ringed PAHs had been degraded, whereas in the autoclaved controls these molecules remained (Paper III: Figure 2). Naphthalene mineralisation and the copy numbers of PAH–RHDα genes were low in the autoclaved controls supporting abiotic reactions explaining the disappearance of total PAHs (Paper III: Figure 3, Table 2).

4.5.3. Concretion bacterial community responses to petroleum hydrocarbons

The gene copy number of PAH–RHDα is supposed to reveal the development of the PAH– degradation potential of a bacterial community under the pressure of aromatic compounds of crude oil (Cebron et al. 2008), although it does not reveal the actual activity (transcription) of the gene. The GP PAH–RHDα gene copy numbers rose only temporarily ten–fold in some of the oxic mixed concretion microcosms. This might be due to high concentration of oil (Rahman et al. 2002) which may have caused toxic effects and bacteria may have favoured more easily degradable crude–oil compounds over PAHs. The significant increase in GN PAH–RHDα gene copy number may reflect the increase in the proportion of Proteobacteria in the concretion and sediment incubations (Paper III: Table 2, Figure 4). Even if this gene codes for the enzyme which requires oxygen, the presence of PAH–degradation genes has also been observed in anoxic petroleum hydrocarbon contaminated subsurface soil (Tuomi et al. 2004). Bacteria may keep the gene in fluctuating oxygen conditions. A total number of 515 OTUs were obtained in the 11 clone libraries named after the crude–oil experiments (Paper III: Figures 4, S1). The bacterial community remained rich even after several weeks of crude–oil exposure, which is in contrast to many previous studies, where bacterial diversity decreases and shifts towards oil–degrading species, such as Thalassolituus oleivorans, Cycloclasticus and Alcanivorax (Kasai et al. 2002, Röling et al. 2002, reviewed by Head et al. 2006 McKew et al. 2007, reviewed by Yakimow et al. 2007). Still Thalassolituus oleivorans has been easily enriched from the concretions in Fe– oxidizing conditions (Paper I: Table 4). Sequences classified as Alphaproteobacteria after oxic incubation of mixed concretions had closest matches found from polluted soil (Nielsen et al. 2014), and Mn–oxide minerals (Miller et al. 2012), conditions possibly resembling the crude–oil exposure experiments. Betaproteobacteria obtained from anoxic incubations of mixed concretions could be classified as Comamonadaceae, a family which includes the known Fe–reducing bacteria classified as Albidoferax (Ramana and Sasikala 2009). Few clones from oxic LL4A samples were affiliated with cultured species Geopsychrobacter electrodiphilus (Deltaproteobacteria) (Holmes et al. 2004), which are known to reduce Fe and to transfer electrons to solid stage through pili. The maintenance of different bacterial communities in concretions and sediments indicated that various communities are capable to tolerate and degrade the compounds of crude oil. This supports the finding that diverse bacterial community rather than enrichment of specific OTUs can

44 lead to more effective degradation of petroleum hydrocarbons in contaminated sediments (Dell’Anno et al. 2012).

Conclusions

x Bacterial community structure in the concretions of the GoF are characterised by high proportions of unclassified bacteria and Proteobacteria. The overall diversity is high. Bacteria colonise both the surface and the interior of a concretion and the enrichment of taxa may result from chemical gradients within the concretion. The cloned sequences had their closest matches found from concretions themselves and other metal–rich environments, yet the concretions are not dominated by the known Fe and Mn –oxidising or – reducing species. (Paper I, III) x Bacteria derived from the concretions are able to oxidise Mn and participate in the formation of Fe–oxides. Therefore, formation of concretions is partly biotic. Different species were enriched depending on the reduced metal and addition of C to the culture media. (Paper I) x Bacteria in the concretions are able to reduce Fe and Mn only with added labile C, in anoxic conditions. Due to the bacterial reduction, Fe and Mn are dissolved and Fe–associated P and As are released to the ambient water. Concretions are significant Fe, Mn, P and As reserve, of which only minor part may be released during prolonged anoxic periods in the GoF. Thus, concretions tolerate anoxia relatively well. (Paper II) x Bacteria hosted by concretions and sediment have similar capacity to degrade crude oil in oxic and anoxic conditions. A part of the disappearance of the crude oil may be explained by attachment to the concretion material and sediment particles. Concretion and sediment bacteria are also able to mineralise naphthalene. The connection between Fe and Mn–reducing bacteria and crude–oil degradation could be shown in few incubations. (Paper III) x The bacterial communities of concretions and sediment differ from one another, before and after the crude–oil exposure. Concretion bacteria are more similar to other bacteria found from metal–rich environments and concretions than sediment bacteria, which in turn are similar to those previously found from sediments. Concretions maintain ecotypes rather than certain species proving that concretions are selective microbial habitats. (Paper III)

Future prospects

Concretions are active players in the sea bottom ecosystem as discussed earlier. This study found indication of concretions contributing to N cycling as well, but further direct studies 2- - about other microbial processes (SO4 reduction, NO3 reduction etc.) will complete our geochemical understanding. Comparative studies on metal and nutrient cycling in the different parts of the Baltic Sea are imperative. The present study focused on Bacteria, leaving the role of Archaea in the concretions aside, even if they occupy nearly every habitat where bacteria are found, and also archaea participate in many environmentally important biogeochemical cycles. Next generation sequencing would create large data set of bacterial and archaeal sequences which would provide more comprehensive picture of microbial communities in concretions. Methods in transcriptomics, single cell isolation and whole genome sequencing will offer detailed information about enriched and isolated species from the concretions. Concretions have intrigued commercial interests as sources of ore and possibly rare elements, but the effects of their large scale mining on the environment or concretions as targets of protection have not yet been assessed.

Acknowledgements This work was carried out in the Marine Research Centre of Finnish Environment Institute. The work was financially supported by Maj and Tor Nessling foundation, K.H. Renlund foundation, University of Helsinki and Maa– ja vesitekniikan tuki ry. I cordially thank my supervisors Kirsten Jørgensen and Jouni Lehtoranta for the opportunity to work with concretion bacteria. You both are wise and patient persons! Thank you for taking me to great conferences, courses and cruises, and for hours of conversations about theoretical and practical aspects of my work. High five to my co–authors Timo Sara–aho, Jari Nuutinen and Anna Reunamo! I sincerely thank my counselling group Susanna Hietanen, Petri Auvinen and Juha Karhu for their valuable support and advice. I extend my gratitude to excellent reviewers Sven Blomqvist and Marko Virta for their useful comments and suggestions to improve the quality of my thesis. Thank you Harri Kankaanpää, Aarno Kotilainen and Joonas Virtasalo for your co– operation, company at conferences and the cruise in the GoF, and giving meaning to “geomicrobiology”. During my efforts to complete this thesis I have received aid and support from several experienced researchers, and friendly colleagues: Lotta Björklöf, Anna Downie, Petri Ekholm, Kaisa Heinonen, Ilse Heiskanen, Sami Huhtala, Laura Häkkinen, Helena Jäntti, Anne Markkanen, Sinikka Pahkala (in memoriam), Jukka Rinkinen, Suvi Tyynelä, Anssi Vähätalo, Yi–Hua Xiao, and wonderful researcher colleagues at the great networks NENuN, FIMIN and EnSTe. I would not have survived without “a PhD therapy group” at SYKE: Aura Nousiainen, Päivi Munne, Noora Perkola, Anna Soirinsuo, Raisa Turja and Pia Välitalo. Thank you Anna–Stiina Heiskanen, Pasi Laiho, Kari Lehtonen and Mari Walls for your firm support as leaders during these years in MRC. I would also like to acknowledge the personnel in the laboratory of Finnish environment institute with gratitude for your work in analyses and supporting the facilities. Special thanks are acknowledged to the crews of r/v Aranda and r/v Muikku, and also the Russian researchers who participated the cruise with me in the GoF. My dear friends Sari, Johanna, Mervi, Anna–Kaisa, Varpu, Laura, Maija, Petra, Riikka… thank you for your unconditional support and answering my phone calls. My parents, thank you for your countless efforts. Dear Markus, thank you for your tolerance and loving companionship during this work. For my self Better done than perfect!

References

Acosta–Gonzáles A, Rosselló–Móra R and Marqués S. 2013. Characterization of the anaerobic microbial community in oil–polluted subtidal sediments: aromatic biodegradation potential after the Prestige oil spill. Environ Microbiol 15: 77–92.

Aeckersberg F, Bak F and Widdel F. 1991. Anaerobic oxidation of saturated hydrocarbons to CO2 by a new type of sulfate–reducing bacterium. Arch Microbiol 1561: 5–14.

Ahmann D, Krumholz LR, Hemond HF, Lovley DR and Morel FMM. 1997. Microbial mobilization of arsenic from sediments of the Aberjona Watershed. Environ Sci Technol 31: 2923–2930.

Akerman NH, Price RE, Pichler T and Amend JP. 2011. Energy sources for chemolithotrophs in an arsenic– and iron–rich shallow–sea hydrothermal system. Geobiology 9: 436–445.

Alain K, Harder J, Widdel F and Zengler K. 2012. Anaerobic utilization of toluene by marine alpha– and gammaproteobacteria reducing nitrate. Microbiology 158: 2946–2957.

Altermann, E. 2014. Invited commentary: lubricating the rusty wheel, new insights into iron oxidizing bacteria through comparative genomics. Front Microbiol 5, 386: 1–4. doi:10.3389/fmicb.2014.00386

Amann R, Fuchs BM and Behrens S. 1995. The identification of microorganisms by fluorescence in situ hybridisation. Curr Opin Biotech 12: 231–236.

Anderson, CR and Cook GM. 2004. Isolation and characterization of arsenate–reducing bacteria from arsenic–contaminated sites in New Zealand. Curr Microbiol 48: 341–347.

Anderson CR, James RE, Fru EC, Kennedy CB and Pedersen K. 2006. In situ ecological development of a bacteriogenic iron oxide–producing microbial community from a subsurface granitic rock environment. Geobiology 4: 29–42.

Anderson RT and Lovley DR. 1999. Naphthalene and benzene degradation under FeIII– reducing conditions in petroleum–contaminated aquifers. Bioremediat J 3: 121–135.

Anufriev GS, Blinov LN, Boltenkov BS and Arif M. 2005. Chemical and isotope composition of Baltic iron–manganese concretions. Tech Phys 50: 663–665.

Archibald FS and Fridovich I. 1981. Manganese and defenses against oxygen toxicity in Lactobacillus plantarum. J Bacteriol 145: 1442–451. 48

Ball P. 2007. Bacteria may be wiring up the soil. Nature 449: 388.

Bamforth SM and Singleton I. 2005. Bioremediation of polycyclic aromatic hydrocarbons: current knowledge and future directions. J Chem Technol Biot 807: 723–736.

Baturin GN. 2009. Geochemistry of ferromanganese nodules in the Gulf of Finland, Baltic Sea. Lithol Miner Resour 44: 411–426.

Baturin GN and Dubinchuk VT. 2009. Composition of ferromanganese nodules from Riga Bay (Baltic Sea). Oceanology 49: 111–120.

Bird LJ, Bonnefoy V and Newman DK. 2011. Bioenergetic challenges of microbial iron metabolism. Trends Microbiol 19: 330–340.

Bissen M and Frimmel FH. 2003. Arsenic – a review. Part I: Occurrence, toxicity, speciation, mobility. Acta Hydrochim Hydrobiol 31: 9–18.

Blöthe M, Wegorzewski A, Müller C, Simon F, Kuhn T and Schippers A. 2015. Manganese–cycling microbial communities inside deep–sea manganese nodules. Environ Sci Technol 49: 7692−7700.

Brakstad OG and Lodeng AGG. 2005. Microbial diversity during biodegradation of crude oil in seawater from the North Sea. Microb Ecol 491: 94–103.

Brettar I, Christen R and Hofle MG. 2002. Rheinheimera baltica gen. nov., sp. nov., a blue–coloured bacterium isolated from the central Baltic Sea. Int J Syst Evol Microbiol 52: 1851–1857.

Bruun A–M, Finster K, Gunnlaugsson HP, Nørnberg P and Friedrich MW. 2010. A comprehensive investigation on iron cycling in a freshwater seep including microscopy, cultivation and molecular community analysis. Geomicrobiol J 27: 15–34.

Burton ED, Johnston SG and Kocar BD. 2014. Arsenic mobility during flooding of contaminated soil: the effect of microbial sulfate reduction. Environ Sci Technol 48: 13660–13667.

Canfield DE, Thamdrup B and Jensen JW. 1993. The anaerobic degradation of organic matter in Danish coastal sediments: iron reduction, manganese reduction, and sulfate reduction. Geochim Cosmochim Ac 57: 3867–3883.

Cappello S, Denaro R, Genovese M, Giuliano L and Yakimov MM. 2007. Predominant growth of Alcanivorax during experiments on "oil spill bioremediation" in mesocosms. Microbiol Res 1622: 185–190.

Carmichael MJ, Carmichael SK, Santelli CM, Strom A and Bräuer SL. 2013. Mn(II)– oxidizing bacteria are abundant and environmentally relevant members of ferromanganese deposits in caves of the upper tennessee river basin. Geomicrobiol J 30: 779–800.

Cebron A, Norini MP, Beguiristain T and Leyval C. 2008. Real–Time PCR quantification of PAH–ring hydroxylating dioxygenase PAH–RHD alpha genes from Gram positive and Gram negative bacteria in soil and sediment samples. J Microbiol Meth 732: 148–159.

Chen Z, Kim K–W, Zhu Y–G, McLaren R, Liu F and He J–Z. 2006. Adsorption (AsIII,V) and oxidation (AsIII) of arsenic by pedogenic Fe–Mn nodules. Geoderma 136: 566–572.

Clement BG, Luther GW and Tebo BM. 2009. Rapid, oxygen–dependent microbial Mn(II) oxidation kinetics at sub–micromolar oxygen concentrations in the Black Sea suboxic zone. Geochim Cosmochim Ac 73: 1878–1889.

Coates JD, Anderson RT, Woodward JC, Phillips EJP and Lovley DR. 1996a. Anaerobic hydrocarbon degradation in petroleum–contaminated harbor sediments under sulfate– reducing and artificially imposed iron–reducing conditions. Environ Sci Technol 30: 2784– 2789.

Coates JD, Phillips EJP, Lonergan DJ, Jenter H and Lovley DR. 1996b. Isolation of Geobacter species from diverse sedimentary environments. Appl Environ Microbiol 62: 1531–1536.

Coates JD, Woodward J, Allen J, Philp P and Lovley DR. 1997. Anaerobic degradation of polycyclic aromatic hydrocarbons and alkanes in petroleum–contaminated marine harbor sediments. Appl Environ Microbiol 639: 3589–3593.

Coleman ML, Hedrick DB, Lovley DR, White DC and Pye K. 1993. Reduction of Fe(III) in sediments by sulphate–reducing bacteria. Nature 361: 436–438.

Corstjens P, de Vrind JPM, Westbroek P and de Vrind–de Jong EW. 1992. Enzymatic iron oxidation by Leptothrix discophora: identification of an iron–oxidizing protein. Appl Environ Microbiol 58: 450–454.

Coulon F, McKew BA, Osborn AM, McGenity TJ and Timmis KN. 2007. Effects of temperature and biostimulation on oil–degrading microbial communities in temperate estuarine waters. Environ Microbiol 91: 177–186. 50

Cummings DE, Snoeyenbos–West OL, Newby DT, Niggemyer AM, Lovley DR, Achenbach LA and Rosenzweig RF. 2003. Diversity of Geobacteraceae species inhabiting metal–polluted freshwater lake sediments ascertained by 16S rDNA analyses. Microb Ecol 46: 257–269.

Dassonville F and Renault P. 2002. Interactions between microbial processes and geochemical transformations under anaerobic conditions: a review. Agronomie 22: 51–68.

Dell'Anno A, Beolchini F, Gabellini M, Rocchetti L, Pusceddu A and Danovaro R. 2009. Bioremediation of petroleum hydrocarbons in anoxic marine sediments: Consequences on the speciation of heavy metals. Mar Pollut Bull 58: 1808–1814.

Dell’Anno A, Beolchini F, Rocchetti L, Luna GM and Danovaro R. 2012. High bacterial biodiversity increases degradation performance of hydrocarbons during bioremediation of contaminated harbor marine sediments. Environ Pollut 167: 85–92.

Diaz RJ and Rosenberg R. 2008. Spreading dead zones and consequences for marine ecosystems. Science 321: 926–929.

2+ Diem D and Stumm W. 1984. Is dissolved Mn being oxidized by O2 in absence of Mn bacteria or surface catalysts? Geochim Cosmochim Ac 48: 1571–1573.

Dong D, Hua X, Li Y, Zhang J and Yan D. 2003. Cd adsorption properties of components in different freshwater surface coatings: the important role of ferromanganese oxides. Environ Sci Technol 37: 4106–4112.

Dong Y, Sanford RA, Locke II RA, Cann IK, Mackie RI and Wouke BW. 2014. Fe–oxide grain coatings support bacterial Fe–reducing metabolisms in 1.7−2.0 km–deep subsurface quartz arenite sand stone reservoirs of the Illinois Basin (USA). Front Microbiol 5, 511: 1– 16. doi: 10.3389/fmicb.2014.00511

Douka CE. 1980. Kinetics of manganese oxidation by cell–free extracts of bacteria isolated from manganese concretions from soil. Appl Environ Microbiol 39: 74–80.

Duan WM, Hedrick DB, Pye K, Coleman ML and White DC. 1996. A preliminary study of the geochemical and microbiological characteristics of modern sedimentary concretions. Limnol Oceanogr 41:1404–1414.

Edlund A, Hardeman F, Jansson JK and Sjöling S. 2008. Active bacterial community structure along vertical redox gradients in Baltic Sea sediment. Environ Microbiol 10: 2051–2063. 51

Edwards, KJ, Bach W and McCollom TM. 2005. Geomicrobiology in oceanography: microbe–mineral interactions at and below the seafloor. Trends Microbiol 13: 449–456.

Edwards KJ, Bach W, McCollom TM and Rogers DR. 2004. Neutrophilic iron–oxidizing bacteria in the ocean: their habitats, diversity, and roles in mineral deposition, rock alteration, and biomass production in the deep–sea. Geomicrobiol J 21: 393–404.

Edwards KJ, Bach W and Rogers DR. 2003a. Geomicrobiology of the ocean crust: a role for chemoautotrophic Fe–bacteria. Biol Bull 204: 180–185.

Edwards KJ, Rogers DR, Wirsen CO and McCollom TM. 2003b. Isolation and characterization of novel psychrophilic, neutrophilic, Fe–oxidizing, chemolithoautotrophic alpha– and gamma–proteobacteria from the deep sea. Appl Environ Microbiol 69: 2906– 2913.

Ehrlich HL. 1963. Bacteriology of manganese nodules. I. Bacterial action on manganese in nodule enrichments. Appl Microbiol 11: 15–19.

Ehrlich HL. 1968. Bacteriology of manganese nodules II. Manganese oxidation by cell– free extract from a manganese nodule bacterium. Appl Microbiol 16: 197–202.

Elmgren R and Larsson U. 2001. Nitrogen and the Baltic Sea: managing nitrogen in relation to phosphorus. The Scientific World 1(S2): 371–377. DOI 10.1100/tsw.2001.291

Emelyanov EM and Kravtsov VA. 2007. Cause of elevated As concentrations in the Baltic Sea and Vistula Lagoon. Geochem Int 45: 798–815.

Emerson D. 2004. Introduction. Geomicrobiol J 21: 369–369.

Emerson D, Field EK, Chertkov O, Davenport KW, Goodwin L, Munk C, Nolan C, Woyke T. 2013. Comparative genomics of freshwater Fe–oxidizing bacteria: implications for physiology, ecology, and systematics. Front Microbiol 4, 254: 1–17. doi:10.3389/fmicb.2013.00254

Emerson D, Fleming EJ, McBeth JM. 2010. Iron–oxidizing bacteria: an environmental and genomic perspective. Annu Rev Microbiol 64: 561–83.

Emerson D and Floyd MM. 2005. Enrichment and isolation of iron–oxidizing bacteria at neutral pH. Methods Enzymol 397: 112–123.

Emerson D and Moyer C. 1997. Isolation and characterization of novel iron–oxidizing bacteria that grow at circumneutral pH. Appl Environ Microbiol 63: 4784–4792.

Emerson D and Moyer CL. 2002. Neutrophilic Fe–oxidizing bacteria are abundant at the Loihi seamount hydrothermal vents and play a major role in Fe oxide deposition. Appl Environ Microbiol 68: 3085–3093.

Emerson D, Rentz JA, Lilburn TG, Davis RE, Aldrich H, Chan C and Moyer CL. 2007. A Novel lineage of Proteobacteria involved in formation of marine Fe–oxidizing microbial mat communities. PLOS One 2(8): e667. doi:10.1371/journal.pone.0000667.

Emerson D and Revsbech NP. 1994a. Investigation of an iron –oxidizing microbial mat community located near Aarhus, Denmark: field studies. Appl Environ Microbiol 60: 4022–4031.

Emerson D and Revsbech NP. 1994b. Investigation of an iron –oxidizing microbial mat community located near Aarhus, Denmark: laboratory studies. Appl Environ Microbiol 60: 4032–4038.

Emerson D, Weiss JV and Megonigal JP. 1999. Iron–oxidizing bacteria are associated with ferric hydroxide precipitates (Fe–plaque) on the roots of wetland plants. Appl Environ Microbiol 65: 2758–2761.

Epstein SS. 2013. The phenomenom of microbial uncultivability. Curr Opin Microbiol 16: 636–642.

Eusterhues K, Rennert T, Knicker H, Kogel–Knabner I, Totsche KU and Schwertmann U. 2011. Fractionation of organic matter due to reaction with ferrihydrite: coprecipitation versus adsorption. Environ Sci Technol 45: 527–533.

Farquhar ML, Charnock JM, Livens FR and Vaughan DJ. 2002. Mechanisms of arsenic uptake from aqueous solution by interaction with goethite, lepidocrocite, mackinawite,and pyrite: An X–ray absorption spectroscopy study. Environ Sci Techol 36: 1757–1762.

Fierer N and Lennon JT. 2011. The generation and maintenance of diversity in microbial communities. Am J Bot 98: 439–448.

Finneran KT, Housewright ME and Lovley DR. 2002. Multiple influences of nitrate on uranium solubility during bioremediation of uranium–contaminated subsurface sediments. Environ Microb 4: 510–516.

Ferrarese E, Andreottola G and Oprea IA. 2008. Remediation of PAH–contaminated sediments by chemical oxidation. J Haz Mat 152: 128–139.

Garnaga G, Wyse E, Azemard S, Stankevičius A and de Mora S. 2006. Arsenic in sediments from the southeastern Baltic Sea. Environ Pol 144: 855–861.

Gasparatos D, Tarenidis D, Haisouti C and Oikonomou G. 2005. Microscopic structure of soil Fe–Mn nodules: environmental implication. Environ Chem Lett 2: 175–178.

Glasby GP, Emelyanov EM, Zhamoida VA, Baturin GN, Leipe T, Bahlo R and Bonacker P. 1997. Environments of formation of ferromanganese concretions in the Baltic sea: a critical review. In: Nicholson K, Hein JR, Bühn B and Dasgupta S (eds.) Manganese mineralization: Geochemistry and mineralogy of terrestrial and marine deposits. Geological society Special publication 119: 213–237.

Glasby GP, Stoffers P, Sioulas A, Thijssen T and Friedrich G. 1982. Manganese nodule formation in the Pacific Ocean: a general theory. Geo–Marine Lett 2: 47–53.

Ghiorse WC and Hirsch P. 1982. Isolation and properties of ferromanganese–depositing budding bacteria from Baltic Sea ferromanganese concretions. Appl Environ Microbiol 43: 1464–1472.

Golterman HL. 1995.The role of the iron hydroxide–phosphate–sulphide system in the phosphate exchange between sediments and overlying water. Hydrobiology 297: 43–54.

Golyshina OV, Pivovarova TA, Karavaiko GI, Kondrat’eva TF, Moore ERB, Abraham W–R, Lunsdorf H, Timmis KN, Yakimov MM and Golyshin PN. 2000. Ferroplasma acidiphilum gen. nov., sp. nov., an acidophilic, autotrophic, ferrous–iron–oxidizing, cell– wall–lacking, mesophilic member of the Ferroplasmaceae fam. nov., comprising a distinct lineage of the Archaea. Int J Syst Evol Microbiol 50: 997–1006.

Graff A and Conrad R. 2005. Impact of flooding on soil bacterial communities associated with poplar (Populus sp.) trees. FEMS Microbiol Ecol 53: 401–415.

Granéli E, Wallström K, Larsson U, Granéli W and Elmgren R. 1990. Nutrient limitation of primary production in the Baltic Sea area. Ambio 19: 142–151.

Grigoriev AG, Zhamoida VA, Gruzdov KA, and Krymsky Grigoriev R Sh. 2013. Age and growth rates of ferromanganese concretions from the Gulf of Finland derived from 210Pb measurements. Oceanology 53: 345–351.

Gunnars A and Blomqvist A. 1997. Phosphate exchange across the sediment-water interface when shifting from anoxic to oxic conditions – an experimental comparison of freshwater and brackish–marine systems. Biogeochemistry 37: 203–226.

Gunnars A, Blomqvist S, Johansson P and Andersson C. 2002. Formation of Fe(III) oxyhydroxide colloids in freshwater and brackish seawater, with incorporation of phosphate and calcium. Geochim Cosmochim Ac 66: 745–758.

Gutierrez T, Singleton DR, Berry D, Yang T, Aitken MD and Teske A. 2013. Hydrocarbon–degrading bacteria enriched by the Deepwater Horizon oil spill identified by cultivation and DNA–SIP. ISME J 7: 2091–2104.

Haahti H and Kangas P. 2006. State of the Gulf of Finland in 2004. MERI –Report series of the Finnish Institute of Marine Research. No 55. Helsinki, Finland.

Hall–Stoodley L, Costerton JW and Stoodley P. 2004. Bacterial biofilms: From the natural environment to infectious diseases. Nature Rev Microbiol 2: 95–108.

Hamilton–Taylor J, Davison W and Morfett K. 1996. A laboratory study of the biogeochemical cycling of Fe, Mn, Zn and Cu across the sediment–water interface of a productive lake. Aquatic Sci 58: 191–209.

Hanert HH. 1992. The genus Gallionella. In: Balows A, Trüper HG, Dworkin M, Harder W, Schleifer K–H (eds.) The Prokaryotes, a Handbook on the Biology of bacteria: ecophysiology, isolation, identification, applications. 2nd ed. Vol. 4. Springer–Verlag. pages 4082–4088.

Hansson M, Andersson L and Axe P. 2011. Areal extent and volume of anoxia and hypoxia in the Baltic Sea, 1960–2011. SMHI Swedish Meteorological and Hydrological Institute, Report Oceanography No. 42, Norrköping, Sweden, 76 pp.

Hao C, Zhang L, Wang L, Li S and Dong H. 2012. Microbial community composition in acid mine drainage lake of Xiang mountain sulfide mine in Anhui province, China. Geomicrobiol J 29: 886–895.

Hauck S, Benz M, Brune A and Schink B. 2001. Ferrous iron oxidation by denitrifying bacteria in profundal sediments of a deep lake (Lake Constance). FEMS Microbiol Ecol 37: 127–134.

Hazen TC, Dubinsky EA, DeSantis TZ, Andersen GL, Piceno YM, Singh N, Jansson JK, Probst A, Borglin SE, Fortney JL, Stringfellow WT, Markus Bill M, Conrad ME, Tom LM, Chavarria KL, Alusi TR, Lamendella R, Joyner DC, Spier C, Baelum J, Auer M, 55

Zemla ML, Chakraborty R, Sonnenthal EL, D’haeseleer P, Holman H–YN, Osman S, Lu Z, Van Nostrand JD, Deng Y, Zhou J, Mason OU. 2010. Deep–Sea oil plume enriches indigenous oil–degrading bacteria. Science 330: 204–208.

He J, Zhang L, Jin S and Zhu Y. 2008. Bacterial communities inside and surrounding soil iron–manganese nodules. Geomicrobiol J 25: 14–24.

Head IM, Jones DM and Roling WFM. 2006. Marine microorganisms make a meal of oil. Nat Rev Microbiol 43: 173–182.

Hedlund BP, Gosink JJ and Staley JT. 1997. Verrucomicrobia div. nov., a new division of the Bacteria containing three new species of Prosthecobacter. Antonie von Leeuwenhoek 72: 29–38.

Hedrich S, Schlömann M and Johnson DB. 2011. The iron–oxidizing proteobacteria. Microbiology 157: 1551–1564.

Heid CA, Stevens J, Livak KJ and Williams PM. 1996. Real time quantitative PCR. Gen Res 6: 986–994.

Helcom 2012. Clean Seas Guide, The Baltic Sea Area, MARPOL 73/78 Special Area. In Stankiewicz M. ed. Katajanokankaari 6 B FI–00160 Helsinki, Finland.

Herlemann DPR, Labrenz M, Jürgens K, Bertilsson S, Waniek JJ and Andersson AF. 2011. Transitions in bacterial communities along the 2000 km salinity gradient of the Baltic Sea. ISME J 5: 1571–1579.

Hohmann C, Winkler E, Morin G and Kappler A. 2010. Anaerobic Fe(II)–oxidizing bacteria show As resistance and immobilize As during Fe(III) mineral precipitation. Environ Sci Technol 44: 94–101.

Holmes DE, Nicoll JS, Bond DR and Lovley DR. 2004. Potential role of a novel psychrotolerant member of the family Geobacteraceae, Geopsychrobacter electrodiphilus gen. nov., sp. nov., in electricity production by a marine sediment fuel cell. Appl Environ Microbiol 70: 6023–6030.

Hugenholtz P, Goebel BM and Pace NR. 1998. Impact of culture–independent studies on the emerging phylogenetic view of bacterial diversity. J Bacteriol 180: 4765–4774.

Hughes MF. 2002. Arsenic toxicity and potential mechanisms of action. Toxicol Lett 133: 1–16. 56

Ingri J. 1985. Geochemistry of ferromanganese concretions and associated sediments in the Gulf of Bothnia. PhD thesis, Luleå university of technology, Sweden. ISSN 0348–8373.

Jensen HS, Mortensen PB, Andersen FØ, Rasmussen E and Jensen A. 1995. Phosphorus cycling in a coastal marine sediment, Aarhus Bay, Denmark. Limnol Oceanogr 40: 908– 917.

Jensen MM, Thamdrup B, Rysgaard S, Holmer M and Fossing H. 2003. Rates and regulation of microbial iron reduction in sediments of the Baltic–North Sea transition. Biogeochemistry 65: 295–317.

Jilbert T and Slomp CP. 2013. Iron and manganese shuttles control the formation of authigenic phosphorus minerals in the euxinic basins of the Baltic Sea. Geochim Cosmochim Ac 107: 155–169.

Jilbert T, Slomp CP, Gustafsson BG and Boer W. 2011. Beyond the Fe–P–redox connection: preferential regeneration of phosphorus from organic matter as a key control on Baltic Sea nutrient cycles. Biogeoscience 8: 1699–1720.

Jørgensen BB. 1978. A comparison of methods for the quantification of bacterial sulfate reduction in coastal marine sediments I. Measurement with radiotracer techniques. Geomicrobiol J 1: 11–27.

Kappler A, Benz M, Schink B and Brune A. 2004. Electron shuttling via humic acids in microbial iron(III) reduction in a freshwater sediment. FEMS Microbiol Ecol 47: 85–92.

Kasai Y, Kishira H and Harayama S. 2002. Bacteria belonging to the genus Cycloclasticus play a primary role in the degradation of aromatic hydrocarbons released in a marine environment. Appl Environ Microbiol 68: 5625–5633.

Knoblauch C and Jørgensen BB. 1999. Effect of temperature on sulphate reduction, growth rate and growth yield in five psychrophilic sulphate–reducing bacteria from Arctic sediments. Environ Microbiol 1: 457–467.

Konhauser KO, Kappler A and Roden EE. 2011. Iron in microbial metabolism. Elements 7: 89–93.

Kristensen E, Kristiansen KD and Jensen MH. 2003. Temporal behaviour of manganese and iron in a sandy coastal sediment exposed to water column anoxia. Estuaries 26: 690– 699.

Kristiansen KD, Kristensen E and Jensen MH. 2002. The influence of water column hypoxia on the behaviour of manganese and iron in sandy coastal marine sediment. Estuar Coast Shelf Sci 55: 645–654.

Kuosmanen T, Peltola M, Raulio M, Pulliainen M, Laurila T, Selin J–F, Huopalainen H and Salkinoja–Salonen M. 2005. Effect of polarization on manganese biofouling of heat exchanger surfaces. In: Müller–Steinhagen H, Malayeri MR and Watkinson AP (eds.) ECI Symposium Series, vol RP2: Proceedings of 6th International Conference on Heat Exchanger Fouling and Cleaning – Challenges and Opportunities.

Langenhoff AAM, Zehnder AJB and Schraa G. 1996. Behaviour of toluene, benzene and naphthalene under anaerobic conditions in sediment columns. Biodegradation 7: 267–274.

LaRiviere D, Harris BC, Autenrieth RL, Bonner JS and Keith N. 2001. Electron acceptor consumption during petroleum degradation in microcosm experiments. In: Leeson A, Foote EA, Banks MK and Magar VS (eds.) Phytoremediation, wetlands, and sediments: The sixth in situ and on–site bioremediation symposium. 25–32. Battelle Press, U.S.A. ISBN 1–57477–115–9.

Leahy JG and Colwell RR. 1990. Microbial degradation of hydrocarbons in the environment. Microbiol Rev 543: 305–315.

Leduc R, Samson R, Al–Bashir B, Al–Hawari J and Cseh T. 1992. Biotic and abiotic disappearance of four PAH compounds from flooded soil under various redox conditions. Water Sci Technol 26: 51–60.

Lee K and Levy EM. 1989. Enhancement of the natural biodegradation of condensate and crude oil on beaches of Atlantic Canada. International Oil Spill Conference Proceedings: Vol. 1989: 479–486.

Lehtoranta J. 2003. Dynamics of sediment phosphorus in the brackish Gulf of Finland. Monographs of the Boreal Environment research 24. ISBN 952–11–1400–2 (PDF).

Lehtoranta J and Heiskanen A–S. 2003. Dissolved iron:phosphate ratio as an indicator of phosphate release to oxic water of the inner and outer coastal baltic Sea. Hydrobiologia 492: 69–84.

Lehtoranta J, Heiskanen A–S and Pitkänen H. 2004. Particulate N and P characterizing the fate of nutrients along the estuarine gradient of the River Neva (Baltic Sea). Estuar Coast Shelf S 61: 275–287.

Lehtoranta J and Pitkänen H. 2003. Binding of phosphate in sediment accumulation areas of the eastern Gulf of Finland, Baltic Sea. Hydrobiol 492: 55–67.

Leinonen R, Akhtar R, Birney E, Bower L, Cerdeno–Tarraga A, Cheng Y, Cleland I, Faruque N, Goodgame N, Gibson R, Hoad G, Jang M, Pakseresht N, Plaister S, Radhakrishnan R, Reddy K, Sobhany S, Hoopen PT, Vaughan R, Zalunin V and Cochrane G. 2010. The European Nucleotide Archive. Nucleid Acids Res 23: 1–4.

Leivuori M and Niemistö L. 1995. Sedimentation of trace metals in the Gulf of Bothnia. Chemosphere 31: 3839–3856.

Leivuori M and Vallius H. 1998. A case study of seasonal variation in the chemical compositions of accumulating suspended sediments in the central Gulf of Finland. Chemosphere 36: 2417–2435.

Leppäranta M and Myrberg K. 2009. Physical oceanography of the Baltic Sea. Springer– Praxis books in geophysical sciences. ISBN 978–3–540–79702–9.

Li C–H, Wong Y–S and Tam NF–Y. 2010. Anaerobic biodegradation of polycyclic aromatic hydrocarbons with amendment of ironIII in mangrove sediment slurry. Bioresource Technol 101: 8083–8092.

Li C–H, Wong Y–S and Tam NF–Y. 2011. Effect of MnIV on the biodegradation of polycyclic aromatic hydrocarbons under low–oxygen condition in mangrove sediment slurry. J Hazard Mater 190: 786–793.

Liblik T and Lips U. 2011. Characteristics and variability of the vertical thermohaline structure in the Gulf of Finlanf in summer. Boreal Environ Res 16 (suppl. A ): 73–83.

Liu J, Wang Z, Belchik SM, Edwards MJ, Liu C, Kennedy DW, Merkley ED, Lipton MS, Butt JN, Richardson DJ, Zachara JM, Fredrickson JK, Rosso KM and Liang Shi L. 2012. Identification and characterization of MtoA: A decaheme c–type cytochrome of the neutrophilic Fe(II)–oxidizing bacterium sideroxydans lithotrophicus ES–1. Frontiers Microbiol 3: 1 –11.

Llorens–Mare`s T, Yooseph S, Goll J, Hoffman J, Vila–Costa M, Borrego CM, Dupont CL and Casamayor EO. 2015. Connecting biodiversity and potential functional role in modern euxinic environments by microbial metagenomics. ISME J 9: 1648–1661.

Lonergan DJ, Jenter HL, Coates JD, Phillips EJP, Schmidt TM and Lovley DR. 1996. Phylogenetic analysis of dissimilatory Fe(III)–reducing bacteria. J Bacteriol. 178: 2402– 2408. 59

Lovley D. 1991. Dissimilatory Fe(III) and Mn(IV) reduction. Microbiol Rev 55: 259–287.

Lovley D. 1995. Bioremediation of organic and metal contaminants with dissimilatory metal reduction. J Indust Microbiol 14: 85–93.

Lovley DR. 2006. Bug juice: harvesting electricity with microorganisms. Nature Rev Microbiol 4: 497–508.

Lovley DR and Anderson RT. 2000. Influence of dissimilatory metal reduction on fate of organic and metal contaminants in the subsurface. Hydrogeol J 8 : 77–88.

Lovley DR, Baedecker MJ, Lonergan DJ, Cozzarelli IM, Phillips EJP and Siegel DI. 1989. Oxidation of aromatic contaminants coupled to microbial iron reduction. Nature 339: 297– 300.

Lovley DR and Coates JD. 2000. Novel forms of anaerobic respiration of environmental relevance. Curr Opin Microbiol 3: 252–256.

Lovley DR, Holmes DE and Nevin KP. 2004. Dissimilatory Fe(III) and Mn(IV) reduction. Adv Microb Physiol 49: 219–285.

Lovley DR and Phillips EJP. 1987. Competitive mechanisms for inhibition of sulfate reduction and methane production in the zone of ferric iron reduction in sediments. Appl Environ Microb 53: 2636–2641.

Lovley DR and Phillips EJP. 1988. Novel mode of microbial energy metabolism: organic carbon oxidation coupled to dissimilatory reduction of iron or manganese. Appl Environ Microb 54: 1472–1480.

Lovley DR, Woodward JC and Chapelle FH. 1994. Stimulated anoxic biodegradation of aromatic hydrocarbons using FeIII ligands. Nature 370: 128–131.

Lozupone CA and Knight R. 2007. Global patterns in bacterial diversity. Proc Nat Acad Sci 104: 11436–11440.

Lu H, Chandran K and Stensel D. 2014. Microbial ecology of denitrification in biological waste water treatment. Water Res 64: 237 –m 254.

Lukkari K, Leivuori M and Kotilainen A. 2009a. The chemical character and behaviour of phosphorus in poorly oxygenated sediments from open sea to organic–rich inner bay in the Baltic Sea. Biogeochemistry 96: 25–48.

Lukkari K, Leivuori M, Vallius H and Kotilainen A. 2009b. The chemical character and burial of phosphorus in shallow coastal sediments in the northeastern Baltic Sea. Biogeochemistry 94: 141–162.

Marcus MA, Manceau A and Kersten M. 2004. Mn, Fe, Zn and As speciation in a fast– growing ferromanganese marine nodule. Geochim Cosmochim Ac 68: 3125–3136.

Martin AJ and Pedersen TF. 2002. Seasonal and interannual mobility of arsenic in a lake impacted by metal mining. Environ Sci Technol 36: 1516–1523.

Mason OU, Di Meo–Savoie CA, Van Nostrand JD, Zhou J, Fisk MR and Giovannoni SJ. 2009. Prokaryotic diversity, distribution and insights into their role in biogeochemical cycling in marine basalts. ISME J 3: 231–242.

McBeth JM, Fleming EJ and Emerson D. 2013. The transition from freshwater to marine iron–oxidizing bacterial lineages along a salinity gradient on the Sheepscot River, Maine, USA. Env Microbiol Rep 5: 453–463.

McKew BA, Coulon F, Osborn AM, Timmis KN and McGenity TJ. 2007. Determining the identity and roles of oil–metabolizing marine bacteria from the Thames estuary, UK. Environ Microbiol 91: 165–176.

McKnight DM, Bencala KE, Zellweger GW, Aiken GR, Feder GL and Thorn KA. 1992. Sorption of dissolved organic carbon by hydrous aluminum and iron oxides occurring at the confluence of Deer Creek with the Snake River, Summit County, Colorado. Environ Sci Technol 26: 1388–1396.

Miller AZ, Dionísio A, Sequeira Braga MA, Hernández–Mariné M, Afonso MJ, Muralha VSF, Herrera LK, Raabe J, Fernandez–Cortes A, Cuezva S, Hermosin B, Sanchez–Moral S, Chaminé H and Saiz–Jimenez C. 2012. Biogenic Mn oxide minerals coating in a subsurface granite environment. Chem Geol 322–323: 181–191.

Morales SE, Cosart TF, Johnson JV and Holben WE. 2009. Extensive phylogenetic analysis of a soil bacterial community illustrates extreme taxon evenness and the effects of amplicon length, degree of coverage, and DNA fractionation on classification and ecological parameters. Appl Environ Microbiol 75: 668–675.

Mort HP, Slomp CP, Gustafsson BG and Andersen TJ. 2010. Phosphorus recycling and burial in Baltic Sea sediments with contrasting redox conditions. Geochim Cosmochim Ac 74: 1350–1362.

Moter A and Göbel GB. 2000. Fluorescence in situ hybridization (FISH) for direct visualization of microorganisms. J Microbiol Meth 41: 85–112.

Mullen MD, Wolf DC, Ferris FG, Beveridge TJ, Flemming CA and Bailey GW. 1989. Bacterial sorption of heavy metals. Appl Environ Microbiol 55: 3143–3149.

Myers CR and Nealson KH. 1988. Bacterial manganese reduction and growth with manganese oxide as the sole electron acceptor. Science 240: 1319–1321.

Nealson KH, Belz A and McKee B. 2002. Breathing metals as a way of life: geobiology in action. Antonie van Leeuwenhoek 81: 215–222.

Neidhardt H, Berner ZA, Freikowski D, Biswas A, Majumder S, Winter J, Gallert C, Chatterjee D and Norra S. 2014. Organic carbon induced mobilization of iron and manganese in a West Bengal aquifer and the muted response of groundwater arsenic concentrations. Chem Geol 367: 51–62.

Nemergut DR, Martin AP and Schmidt SK. 2004. Integron diversity in heavy–metal– contaminated mine tailings and inferences about integron evolution. Appl Environ Microbiol 70: 1160–1168.

Nielsen LP. 1992. Denitrification in sediment determined from nitrogen isotope pairing. FEMS Microbiol Ecol 86: 357–362.

Nielsen LP, Risgaard–Petersen N, Fossing H, Christensen PB and Sayama M. 2010. Electric currents couple spatially separated biogeochemical processes in marine sediment. Nature 463: 1071–1074.

Nielsen MB, Kjeldsen KU, Lever MA and Ingvorsen K. 2014. Survival of prokaryotes in a polluted waste dump during remediation by alkaline hydrolysis. Ecotoxicol 233: 404–418.

Ofek M, Hadar Y and Minz D. 2009. Comparison of effects of compost amendment and of single–strain inoculation on root bacterial communities of young cucumber seedlings. Appl Environ Microbiol 75: 6441–6450.

Omoregie EO, MastalerzV, de Lange G, Straub KL, Kappler A, Roy H, Stadnitskaia A, Foucher JP and Boetius A. 2008. Biogeochemistry and community composition of iron– 62 and sulfur–precipitating microbial mats at the Chefren mud volcano Nile Deep Sea Fan, Eastern Mediterranean. Appl Environ Microb 74: 3198–3215.

Oscarsson DW, Huang PM and Hammer VT. 1983. Oxidation and sorption of arsenite by manganese dioxide as influenced by surface coatings of iron and aluminium oxides and calcium carbonate. Water Air Soil Poll 20: 233–244.

Paisse S, Goni–Urriza M, Coulon F and Duran R. 2010. How a bacterial community originating from a contaminated coastal sediment responds to an oil input. Microb Ecol 60: 394–405.

Pakhomova SV, Hall POJ, Kononets MY, Rozanov AG, Tengberg A and Vershinin AV. 2007. Fluxes of iron and manganese across the sediment–water interface under various redox conditions. Mar Chem 197: 319–331.

Park KS, Sims RC, Dupont RR, Doucette WJ and Matthews JE. 1990. Fate of PAH compounds in two soil types: Influence of volatilization, abiotic loss and biological activity. Environ Toxicol Chem 9: 187–195.

Pelletier E, Delille D and Delille B. 2004. Crude oil bioremediation in sub–Antarctic intertidal sediments: chemistry and toxicity of oiled residues. Mar Environ Res 57: 311– 327.

Pfeffer C, Larsen S, Song J, Dong M, Besenbacher F, Meyer RL, Kjeldsen KU, Schreiber L, Gorby YA, El–Naggar MY, Leung KM, Schramm A, Risgaard–Petersen N and Nielsen LP. 2012. Filamentous bacteria transport electrons over centimetre distances. Nature 1–4. doi:10.1038/nature11586

Pitkänen H. 1994. Eutrophication of the Finnish coastal waters: Origin, fate and effects of riverine nutrient fluxes. Publications of the Water and Environment Research Institute. National Board of Waters and the Environment, Finland. No. 18. 45p. ISBN 951–47– 9857–0.

Pitkänen H, Lehtoranta J, Peltonen H, Laine A, Kotta J, Kotta I, Moskalento P, Mäkinen A, Kangas P, Perttilä M and Kiirikki M. 2003. Benthic release of phosphorus and its relation to environmental conditions in the estuarial Gulf of Finland, Baltic Sea in the early 2000s. Proceedings of the Estonian Academy of Sciences, Biology and Ecology 52: 173– 193.

Pitkänen H, Lehtoranta J and Räike A. 2001. Internal nutrient fluxes counteract decreases in external load: the case of the estuarial eastern Gulf of Finland, Baltic Sea. Ambio 30: 195–201. 63

Postma D and Appelo CAJ. 2000. Reduction of Mn –oxides by ferrous iron in a flow system: column experiment and reactive transport modelling. Geochim Cosmochim Ac 64: 1237–1247.

Raab A and Feldmann J. 2003. Microbial transformation of metals and metalloids. Sci Prog 86: 179–202.

Radajewski S, Ineson P, Parekh NR and Murrell JC. 2000. Stable–isotope probing as a tool in microbial ecology. Nature 403: 646–649.

Rahman KSM, Thahira–Rahman J, Lakshmanaperumalsamy P and Banat IM. 2002. Towards efficient crude oil degradation by a mixed bacterial consortium. Biores Technol 85: 257–261.

Ramana CV and Sasikala C. 2009. Albidoferax, a new genus of Comamonadaceae and reclassification of Rhodoferax ferrireducens (Finneran et al., 2003) as Albidoferax ferrireducens comb. nov. J Gen Appl Microbiol 55: 301–304.

Rasmussen R, Morrison T, Herrman M and Wittwer C. 1998. Quantitative PCR by continuous fluorescence monitoring of a double strand DNA specific binding dye. Biochemica 2: 8–11.

Reguera G, MacCarthy KD, Mehta T, Nicoll JS, Tuominen MT and Lovley DK. 2005. Extracellular electron transfer via microbial nanowires. Nature 435: 1098–1101.

Rentz J, Turner IP and Ullman JL. 2009. Removal of phosphorus from solution using biogenic iron oxides. Water Res. 43: 2029–35.

Riedel T, Zakb D, Biestera H and Dittmar T. 2013. Iron traps terrestrially derived dissolved organic matter at redox interfaces. Proc Nat Ac Sci 110: 10101–10105.

Reunamo A, Riemann L, Leskinen P and Jørgensen KS. 2013. Dominant petroleum hydrocarbon–degrading bacteria in the Archipelago Sea in South–West Finland Baltic Sea belong to different taxonomic groups than hydrocarbon degraders in the oceans. Mar Pollut Bull 72: 174–180.

Roden EE and Edmonds JW. 1997. Phosphate mobilization in iron–rich anaerobic sediments: Microbial Fe(III) oxide reduction versus iron–sulfide formation. Archiv fur Hydrobiologie 39: 347–378.

Roden EE, Sobolev D, Glazer B and Luther GW. 2004 Potential for microscale bacterial Fe redox cycling at the aerobic–anaerobic interface. Geomicrobiol J 21: 379–391.

Roden EE and Wetzel RG. 2003. Competition between Fe(III)–reducing and methanogenic bacteria for acetate in iron–rich freshwater sediments. Microb Ecol 45: 252–258.

Roden EE and Zachara JM. 1996. Microbial reduction of crystalline iron(III) oxides: influence of oxide surface area and potential for cell growth. Environ Sci Technol 30: 1618–1628.

Rona PA. 2003. Resources of the sea floor. Science 299: 673–674.

Rydin E, Malmaeus JM, Karlsson OM and Jonsson P. 2011. Phosphorus release from coastal Baltic Sea sediments as estimated from sediment profiles. Estuar Coast Shelf Sci 92: 111–117.

Röling WFM, Milner MG, Jones DM, Lee K, Daniel F, Swannell RJP and Head IM. 2002. Robust hydrocarbon degradation and dynamics of bacterial communities during nutrient– enhanced oil spill bioremediation. Appl Environ Microbiol 68: 5537–5548.

Santelli CM, Orcutt BN, Banning E, Bach W, Moyer CL, Sogin ML, Staudigel H and Edwards KJ. 2008. Abundance and diversity of microbial life in ocean crust. Nature 453: 653–656.

Schippers A, Neretin LN, Lavik G, Leipe T and Pollehne F. 2005. Manganese(II) oxidation driven by lateral oxygen intrusions in the western Black Sea. Geochim Cosmochim Ac 69: 2241–2252.

Seewald JS. 2003. Organic–inorganic interactions in petroleum–producing sedimentary basins. Nature 426: 327–333.

Sharma P, Ofner J and Kappler A. 2010. Formation of binary and ternary colloids and dissolved complexes of organic matter, Fe and As. Environ Sci Technol 44: 4479–4485.

Shokralla S, Spall JL, Gibson JF and Hajibabaei M. 2012. Next–generation sequencing technologies for environmental DNA research. Mol Ecol 21: 1794–1805.

Singer E, Heidelberg JF, Dhillon A and Edwards KJ. 2013. Metagenomic insights into the dominant Fe(II) oxidizing Zetaproteobacteria from an iron mat at Lo’ihi, Hawai’I. Front Microbiol 4, 52: 1–9. doi: 10.3389/fmicb.2013.00052

Slomp CP, Mort HP, Jilbert T, Reed DC, Gustafsson BG and Wolthers M. 2013. Coupled dynamics of iron and phosphorus in sediments of an oligotrophic coastal basin and the impact of anaerobic oxidation of methane. PLoS ONE 8: e62386. doi:10.1371/journal.pone.0062386

Spiro TG, Bargar JR, Sposito G and Tebo BM. 2008. Bacteriogenic manganese oxides. Acc Chem Res 43: 2–9.

Stein LY, La Duc MT, Grundl TJ and Nealson KH. 2001. Bacterial and archaeal populations associated with freshwater ferromanganous micronodules and sediments. Environ Microbiol 3: 10–8.

Steingruber AM, Friedrich J, Gächter R and Wehrli B. 2001. Measurement of denitrification in sediments with the 15N isotope pairing technique. Appl Environ Microbiol 67: 3771–3778.

Straub KL, Benz M, Schink B and Widdel F. 1996. Anaerobic, nitrate–dependent microbial oxidation of ferrous iron. Appl Environ Microbiol 62: 1458–1460.

Straub KL, Benz M, Schink B and Widdel F. 2001. Iron metabolism in anoxic environments at near neutral pH. FEMS Microbiol Ecol 34: 181–186.

Straub KL and Buchholz –Cleven BEE. 1998. Enumeration and detection of anaerobic ferrous iron– oxidizing, nitrate–reducing bacteria from diverse European sediments. Appl Environ Microbiol 6: 4846–4856.

Sun W, Sierra–Alvarez R, Milner L, Oremland R and Field JA. 2009. Arsenite and ferrous iron oxidation linked to chemolithotrophic denitrification for the immobilization of arsenic in anoxic environments. Environ Sci Technol 43: 6585–6591.

Sunda WG and Kieber DJ. 1994. Oxidation of humic substances by manganese oxides yields low–molecular–weight organic substrates. Nature 367: 62–64.

Surdam RC, Jiao ZS and MacGowan DB. 1993. Redox reactions involving hydrocarbons and mineral oxidants: A mechanism for significant porosity enhancement in sandstones. Bull Am Assoc Petrol Geol 77: 1509–1518.

Tebo BM, Bargar JR, Clement BG, Dick GJ, Murray KJ, Parker D, Verity R and Webb SM. 2004. Biogenic manganese oxides: Properties and mechanisms of formation. Annu Rev Earth Planet Sci 32: 287–328.

Tebo BM and Emerson S. 1986. Microbial manganese(II) oxidation in the marine environment: a quantitative study. Biogeochemistry 2: 149–161.

Tebo BM, Johnson HA, McCarthy JK and Templeton AS. 2005. Geomicrobiology of manganese(II) oxidation. Trends Microbiol 13:421–428.

Tebo BM and Villalobos M. 2005. Introduction: Advances in the geomicrobiology and biogeochemistry of manganese and iron oxidation. Geomicrobiol J 22: 77–78.

Temple KL and Colmer AR. 1951. The autotrophic oxidation of iron by a new bacterium: Thiobacillus Ferrooxidans. J Bacteriol 62: 605–611.

Templeton AS, Staudigel H and Tebo BM. 2005. Diverse Mn(II) –oxidizing bacteria isolated from submarine basalts at Loihi seamount. Geomicrobiol J 22: 127–139.

Thamdrup B, Glud RN and Hansen JW. 1994. Manganese oxidation and in situ manganese fluxes from a coastal sediment. Geochim Cosmochim Ac 58: 2563–2570.

Thamdrup B, Rossello–Mora R and Amann R. 2000. Microbial manganese and sulfate reduction in Black Sea shelf sediments. Appl Environ Microbiol 66: 2888–2897.

Tully BJ and Heidelberg JF. 2013. Microbial communities associated with ferromanganese nodules and the surrounding sediments. Front Microb 4: 1–10.

Tuomi PM, Salminen JM and Jørgensen KS. 2004. The abundance of nahAc genes correlates with the 14C–naphthalene mineralization potential in petroleum hydrocarbon– contaminated oxic soil layers. FEMS Microb Ecol 51: 99–107.

Tuominen L, Kairesalo T and Hartikainen H. 1994. Comparison of methods for inhibiting bacterial activity in sediment. Appl Environ Microbiol 60: 3454–3457.

Vahtera E, Autio R, Kaartokallio H and Laamanen M. 2010. Phosphate addition to phosphorus–deficient Baltic Sea plankton communities benefits nitrogen–fixing Cyanobacteria. Aquat Microb Ecol 60:43–57.

Vandieken V, Finke N and Jørgensen BB. 2006. Pathways of carbon oxidation in an Arctic fjord sediment (Svalbard) and isolation of psychrophilic and psychrotolerant Fe(III)– reducing bacteria. Mar Ecol Prog Ser 322: 29–41.

Vandieken V, Finke N and Thamdrup B. 2014. Hydrogen, acetate, and lactate as electron donors for microbial manganese reduction in a manganese–rich coastal marine sediment. FEMS Microbiol Ecol 87:733–745.

Vandieken V, Pester M, Finke N, Hyun JH, Friedrich MW, Loy A and Thamdrup B. 2012. Three manganese oxide–rich marine sediments harbor similar communities of acetate– oxidizing manganese–reducing bacteria. ISME J 6: 2078–2090.

Vallius H. 2012. Arsenic and heavy metal distribution in the bottom sediments of the Gulf of Finland through the last decades. Baltica 25: 23–32.

Vallius H, Zhamoida V, Kotilainen A and Ryabchuk D. 2011. Seafloor desertification – a future scenario for the Gulf of Finland? Central and Eastern European Development Studies 365–372. DOI: 10.1007/978–3–642–17220–5_17.

Vasconcelos C and McKenzie JA. 2009. The descent of minerals. Science 323: 218–219.

Viggor S, Juhanson J, Joesaar M, Mitt M, Truu J, Vedler E and Heinaru A. 2013. Dynamic changes in the structure of microbial communities in Baltic Sea coastal seawater microcosms modified by crude oil, shale oil or diesel fuel. Microbiol Res 168: 415–427.

Väli G, Meier HEM and Elken J. 2013. Simulated halocline variability in the Baltic Sea and its impact on hypoxia during 1961–2007. J Geophys Res Oceans 118: 6982–7000.

Wang Q, Garrity GM, Tiedje JM and Cole JR. 2007. Naïve Bayesian classifier for rapid assignment of rRNA sequences into the new bacterial taxonomy. Appl Environ Microbiol 73: 5261–7.

Wang X, Schröder HC, Wiens M, Shloßmacher U and Müller WEG. 2009. Manganese/polymetallic nodules: micro–structural characterization of exolithobiontic– and endolithobiontic microbial biofilms by scanning electron microscopy. Micron 40: 350– 358.

Wang Y, Wan R, Zhang SY and Xie SG. 2012. Anthracene biodegradation under nitrate– reducing condition and associated microbial community changes. Biotechnol Bioproc E 17: 371–376.

Weber K, Achenbach LA and Coates JD. 2006. Microorganisms pumping iron: anaerobic microbial iron oxidation and reduction. Nature Rev Microbiol 4: 752–764.

Weiss JV, Rentz JA, Plaia T, Neubauer SC, Merrill–Floyd M, Lilburn T, Bradburne C, Megonigal JP and Emerson D. 2007. Characterization of neutrophilic Fe(II)–oxidizing 68 bacteria isolated from the rhizosphere of wetland plants and description of Ferritrophicum radicicola gen. nov. sp. nov., and Sideroxydans paludicola sp. nov. Geomicrobiol J 24: 559–570.

Whitman WB, Coleman DC and Wiebe WJ. 1998. Prokaryotes: the unseen majority. Proc Nat Acad Sci USA 95: 6578–6583.

Widdel F and Rabus R. 2001. Anaerobic biodegradation of saturated and aromatic hydrocarbons. Curr Opin Biotechnol 12: 259–276.

Winderl C, Anneser B, Griebler C, Meckenstock RU and Lueders T. 2008. Depth–resolved quantification of anaerobic toluene degraders and aquifer microbial community patterns in distinct redox zones of a tar oil contaminant plume. Appl Environ Microbiol 74: 792–801.

Winterhalter B. 1966. Pohjanlahden ja Suomenlahden rautamangaani–saostumista. Geoteknillisiä julkaisuja N:o 69, Geologian tutkimuslaitos, Otaniemi, Finland.

Wu Y–H, Liao L, Wang C–S, Ma W–L, Meng F–X, Wu M and Xu X–W. 2013. A comparison of microbial communities in deep–sea polymetallic nodules and the surrounding sediments in the Pacific Ocean. Deep–Sea Res 79: 40–49.

Yakimov MM, Timmis KN and Golyshin PN. 2007. Obligate oil–degrading marine bacteria. Curr Opin Biotech 18: 257–266.

Yeung CV, Van Stempvoort DR, Spoelstra J, Bickerton G, Voralek J and Greer CW. 2013. Bacterial community evidence for anaerobic degradation of petroleum hydrocarbons in cold climate groundwater. Cold Reg Sci Technol 86: 55–68.

Zhamoida V, Grigoriev A, Gruzdov K and Ryabchuk D (2007) The influence of ferromanganese concretions –forming processes in the eastern Gulf of Finland on the marine environment. In Vallius H, ed. Holocene sedimentary environment and sediment geochemistry of the Eastern Guls of Finland, Baltic Sea. Geological Survey of Finland, Special Paper 45: 21–32. Espoo, Finland.

Zhang F–S, Lin CY, Bian LZ, Glasby GP and Zhamoida VA. 2002. Possible evidence for the biogenic formation of spheroidal ferromanganese concretions from the Eastern Gulf of Finland, Baltic Sea. Baltica 15: 23–29.

Zhang G, Liu F, Liu H, Qu J and Liu R. 2014. Respective role of Fe and Mn oxide contents for arsenic sorption in iron and manganese binary oxide: an X–ray absorption spectroscopy investigation. Environ Sci Technol 48: 10316–10322.

Zhang L–M, Liu F, Tan W–F, Feng X–H, Zhu Y–G and He J. 2008. Microbial DNA extraction and analyses of soil iron–manganese nodules. Soil Biol Biochem 40: 1364– 1369.

Zhang T, Bain TS, Nevin KP, Barlett MA and Lovley DR. 2012. Anaerobic benzene oxidation by Geobacter species. Appl Environ Microbiol 78: 8304–8310.

Zheng G, Selvam A and Wong JWC. 2011. Rapid degradation of lindane (γ– hexachlorocyclohexane) at low temperature by Sphingobium strains. Int Biodeter Biodegr 65: 612–618.

Ziemke F, Hofle MG, Lalucat J and Rossello–Mora R. 1998. Reclassification of Shewanella putrefaciens Owen's genomic group II as Shewanella baltica sp. nov. Int J Syst Bacteriol 48: 179–186.

Geomicrobiology Journal (2014) 31, 263–275 Copyright C Taylor & Francis Group, LLC ISSN: 0149-0451 print / 1521-0529 online DOI: 10.1080/01490451.2013.819050 Iron–Manganese Concretions Sustaining Microbial Life in the Baltic Sea: The Structure of the Bacterial Community and Enrichments in Metal-Oxidizing Conditions

PIRJO YLI-HEMMINKI∗, KIRSTEN S. JØRGENSEN, and JOUNI LEHTORANTA Marine Research Centre, Finnish Environment Institute, Helsinki, Finland

Received January 2013, Accepted June 2013

The abundant deposits of spherical iron-manganese concretions in the Gulf of Finland are colonized by bacteria in vast numbers. Communities on the surface and in the porous interior have formed two separate clusters, in accordance with their genetic differences. The overall bacterial community in the concretions was highly diverse, representing 12 phyla. Half of the bacteria were affiliated with the most common classes of Proteobacteria, while a third of the bacteria were unclassified. Cloned 16S rRNA-gene sequences of the concretion bacteria showed high scores for similarity to the sequences obtained from sea sediments, metal-rich environments, and ocean crust. The clone library of native concretions was not dominated by known Fe- and Mn-oxidizing species. Known Mn-oxidizing bacteria Sphingomonas, Pseudomonas,andBacillus were enriched in experiments with Mn2+-containing liquid media, whereas Prosthecobacter (Verrucomicrobia) and Rheinheimera were enriched in semisolid media possibly better simulating the natural conditions in the concretions. In a corresponding experiment, the Fe2+-oxygen gradient favored the enrichment of Shewanella baltica and Thalassolituus oleivorans, which are known to reduce Fe and to degrade petroleum hydrocarbons, respectively. An individual spherical concretion forms a microcosm for a diverse microbial community having potential to oxidize Fe and Mn as shown in cultivation experiments. Therefore, bacteria may significantly affect the formation of the concretions in the Gulf of Finland. Keywords: Fe-oxidizing bacteria, gradient culture, marine ferromanganese nodules, Mn-oxidizing bacteria

Introduction rate, and geochemical properties, these concretions may have significant influence on the biogeochemical cycles of Fe, Mn, Large deposits of iron-manganese (Fe-Mn) concretions (i.e., P, and harmful metals in the Baltic Sea. marine nodules) are situated on the surface of silty-clayey mud Bacteria are able to oxidize reduced species of Fe and Mn bottoms in the Gulf of Bothnia and the Gulf of Finland, which and thereby precipitate and accumulate metal oxides in the are parts of the Baltic Sea, the world’s largest brackish-water sediment (Tebo and Villalobos 2005). In environments rich in basin. Concretions cover approximately 10% of the seafloor ferrous iron, bacteria can derive energy for growth from Fe (Winterhalter 1966), with the estimates of their density varying oxidation. Among the first of these described were the aci- from 2–15 kg m−2 in the Gulf of Bothnia (Ingri 1985) to dophilic Acidithiobacillus ferrooxidans (previously Thiobacil- 20–50 kg m−2 in the eastern Gulf of Finland (Zhamoida et al. lus ferrooxidans)andLeptospirillum ferrooxidans.Inneutral 2007). The concretions are rich in Fe and Mn, along with co- conditions, bacteria compete with chemical oxidation of Fe precipitated P and As (Baturin and Dubinchuk 2009; Marcus (Konhauser et al. 2011). Environmental geochemical charac- et al. 2004). Growth-rate estimates for the concretions are in teristics such as reduced and oxidized metals in basalts have the 0.03–0.06 mm a−1 range (Zhamoida et al. 2007), which encouraged the search for neutrophilic Fe-oxidizing bacteria is considerably higher than those for oceanic concretions, at (Mason et al. 2009) other than classical Gallionella ferruginea <10 mm / 106 a (Glasby et al. 1982). The large reservoir and Leptothrix ochracea. of Fe and Mn has attracted commercial interest as Mn ore It has been inferred that basaltic Mn, S, and Fe phases may (Anufriev et al. 2005). Because of their abundance, growth supply energy for chemolithoautotrophic growth of microorganisms in deep-sea environment (Santelli et al. 2008). Mi- croaerophilic bacteria are found from oxic-anoxic interfaces ∗Address correspondence to Pirjo Yli-Hemminki, Marine rich in Fe, such as fresh-water seepage (Emerson and Revs- Research Centre, Finnish Environment Institute, P.O. Box bech 1994) and oligotrophic, metalliferous ocean sediments 140, Helsinki FI-00251, Finland; Email: pirjo.yli-hemminki@ (Edwards et al. 2003a, 2003b). Furthermore, bacteria have ymparisto.fi been observed to oxidize Fe under anoxic conditions by using Color versions of one or more of the figures in the article can be nitrate as an electron acceptor (Hauck et al. 2001; Straub et al. found online at www.tandfonline.com/ugmb. 1996). 264 Yli-Hemminki et al.

Mn2+ is not as readily oxidized chemically as Fe2+ is, and prior to sampling confirmed that the areas of the seafloor were bacterial Mn2+ oxidation is responsible for the bulk of the covered by Fe-Mn-concretions at several depths. The samples Mn oxide formation in the sediments (Thamdrup et al. 1994). were stored in plastic bags in near-bottom sea water or artifi- ◦ Biological Mn2+ oxidation is much faster than abiotic Mn2+ cial brackish sea water at 5 C. The water content of 17 con- ◦ oxidation (Tebo and Emerson 1986; Tebo et al. 2004). Within cretions was determined through drying overnight at 105 C. only weeks, bacteria can foul and form Mn micronodules on Furthermore, the outer layer and the interior of concretions the metal surfaces of heat exchangers that use Baltic sea water (nine replicates) were separated with sterile spatulas and stored (Kuosmanen et al. 2005). Microstructures of bacteria and the for further DNA analysis. birnessite and manganosite phase in Fe-Mn-concretions have been reported from the Baltic Sea (Zhang et al. 2002), and electron-microscopy studies have revealed biofilms attached Enumeration of Bacteria by Epifluorescence Microscopy to oceanic nodules (Wang et al. 2009). Bacteria were extracted by homogenization of 10 g wet weight Studies of concretions in soil have shown that the clones of (WW) of solid concretions with 90 mL of 0.9% NaCl-solution, 16S rRNA genes had a close relationship with Mn2+-oxidizing 1 mL of 10.4% Na5O10P3, and 0.1 mL of 2% Tween 80- bacteria in aquatic Fe-Mn-concretions (He et al. 2008; Zhang solution. The sample was treated with formaldehyde to a et al. 2008). Culture-based microbiological studies of concre- final concentration of 0.7% for possible storage. The slurry tions are rare (Ghiorse and Hirsch 1982), although the role was filtered through a 5-μm Millipore membrane filter at- of Mn-oxidizing bacteria in the formation of concretions had tached to a syringe, and the syringe was washed with 1 M already been suggested, for example, by Ehrlich in 1963. KCl, which was also filtered to the sample. The samples The concretions are metal-oxide-rich environments, which were filtered through Whatman 0.2-μm filters and stained may create selective pressure on the formation of the microbial with DAPI (4,6-diamidino-2-phenylindole) (10 μgmL−1)for community. Biofilms in and outside Mn and Fe oxides can be 5 min. dominated by metal-reducing and -oxidizing bacteria (Stein From the liquid experiments, 10-mL samples were taken et al. 2001). It is likely that bacteria play a significant role in and fixed with 200 μL of 37% formaldehyde. The solutions the precipitation of Fe and Mn, but the bacterial community’s were filtered through Whatman mixed cellulose ester 0.2-μm structure in the concretions of the Baltic Sea is not known filters below Whatman cyclopore 0.2-μm polycarbonate filters at present. To understand their formation, there is a need for and stained with DAPI for 5 min. The filters were viewed via knowledge of the microbial communities and the processes a Leica DM RXA2 epifluorescence microscope. Shapes of they maintain in the concretions. In this study, we examined bacteria were noted, and the number of bacteria was counted the bacterial community structure of the concretions and com- within 10 grids. pared the results to those for clones and bacteria isolated from various environments. To investigate the role of bacteria in the growth of concretions, we used multiple culturing methods to DNA Extractions, PCR Technique, Clone Library, test which bacterial groups are enriched in oxidative condi- RFLP Analysis, and Sequencing tions accompanied by reduced Fe and Mn. The DNA was extracted from whole or separated parts of the concretions using the FastDNA SPIN Kit for Soil (Qbiogene; Materials and Methods Carlsbad, CA, USA) and a FastPrep Instrument homogenizer in accordance with the manufacturer’s instructions. The DNA concentration was checked with a Qubit fluorometer (Invitro- Fe-Mn-Concretion Sampling gen Molecular Probes; Carlsbad, CA, USA). Concretions were collected with a van Veen sampler from the The gene encoding the bacterial 16S rRNA was amplified Gulf of Finland, in the Baltic Sea (see Table 1). Echo-sounding with general primers specific for bacteria PRBA338f: 5 AC

Table 1. The sampling sites of iron-manganese concretions from the Gulf of Finland

Water Shape and size Sampling site and year depth (m) of concretions Analysis

2007 Gulf of Finland 59 Spherical Dry matter N60◦ 01.00 and E26◦ 04.85 0.5–2 cm in diameter Clone library Enrichment of Mn-oxidizers in liquid Microscopy 2010 Finnish Estonian EEZ area (st 1) 62 Spherical, Dry matter N59◦ 57.43 and E25◦ 12.76 0.5–2 cm in diameter Enrichment of Fe- and Mn-oxidizers in gradient Microscopy 2010 Finnish Estonian EEZ area (st 12) 54 Spherical, DGGE from the surface and interior of concretions N59◦ 57.24 and E24◦ 59.58 0.5–2 cm in diameter Enrichment of Fe- and Mn-oxidizers in gradient Microscopy Bacterial Community in Marine Fe–Mn Concretions 265

TCC TAC GGG AGG CAG CAG 3 (Øvreas˚ et al. 1997) GCG GCG GGC GGG GCG GGG GCA CGG GGG GAC and DS907r: 5 CCC CGT CAA TTC CTT TGA GTT T 3 TCC TAC GGG AGG CAG CAG 3 and DS907r: 5 CCC (Teske et al. 1996). The PCR reaction-mixture (49 μL) was CGT CAA TTC CTT TGA GTT T 3. The PCR products prepared: 31.5 μL mol-grade water, 5 μL(NH4)2SO4 buffer were loaded onto a 6% polyacrylamide gel in 0.5× TAE (Tris/ (10×), 5 μLMgCl2 (25 mM), 1 μL dNTPs (10 mM each), Acetic Acid/ EDTA). The gel had a denaturing gradient rang- 2 μL primer mix (5 μM), 2 μL Taq polymerase (1 U μL−1) ing from 35% to 80%. The electrophoresis was performed in (Fermentas Int., Inc.; Canada), 2.5 μLBSAand1μLof a vertical V20-HCDCY system (Scie-Plas Ltd., UK) at 60◦C, template were used. The amplification was performed with first for 10 min at 150 V and subsequently for 16 h at 70 V.Af- PTC-200 DNA Engine (MJ Research; Waltham, MA, USA) ter electrophoresis, the gel was stained for 30 min with SYBR as follows: initial denaturing (3 min at 94◦C) followed by 25 Green nucleic acid gel stain (100 mL 0.5X TAE + 10 μLSYBR cycles of denaturation (45 s at 94◦C), primer annealing (45 s Green), rinsed for 30 min in 0.5X TAE, and photographed at 53◦C), and primer extension (1 min at 72◦C). on a blue light transillumination table with Gel Doc 2000. The program was finished with a final extension step (7 min Similarity between sequence profiles was estimated with clus- at 72◦C and end at 15◦C). The PCR products were purified ter analysis using unweighted, average clustering (UPGMA) by the Wizard SV Gel and PCR Clean-Up System (Promega; of the DGGE sequence profiles, analyzed by Quantity One Madison, WI, USA) and checked for the right size (∼569 bp) Quantitation software (Bio-Rad Laboratories; Hercules, CA, by gel electrophoresis on a 2.5% agarose gel (120 V, 60 min). USA). The purified amplification products were cloned using the TOPO TA Cloning Kit for Sequencing version O, with the plasmid vector of expression PCR 4-TOPO (Invitrogen) in Enrichment of Mn-oxidizers in Liquid Culture line with the manufacturer’s instructions. The cloned inserts The composition of the artificial brackish water used as were excised for sequencing, with PCR performed directly growth medium was modified after the work of Tarkiainen from bacterial colonies with vector-specific M13 primers. and Rinne (1973) (see Table 2), and it corresponds to the The restriction enzymes cleave the DNA at specific posi- quality of the water in the central Gulf of Finland. In the tions and allow detection of variation in the DNA sequence, Mn-oxidizing experiment 2-L Erlenmeyer bottles were filled i.e., restriction fragment length polymorphism (RFLP). Clones for the RFLP were selected by picking of random μ colonies. The PCR products (10 L) of the clones were re- Table 2. The composition of artificial brackish sea water used in stricted with 20 U of enzymes Hae III and Msp I (New Eng- enrichment cultures land Biolabs; Beverly, MA, USA). Restriction was performed in 1× NEBuffer 2 solution at 37◦C for 3 h. Concentration Digested fragments were separated by electrophoresis −1 μ −1 (120 V, 110 min) in 0.5× TBE (Tris/Borate/EDTA) buffer Component (mg L )(gL ) using 4% MetaPhor agarose gel (Cambrex Bio Science, Inc.; Salts Walkersville, MD, USA) in accordance with manufacturer’s NaCl 4680 instructions. The gels were stained with SYBR Safe for 30 min MgCl2 1360 and photographed under blue light. Different restriction pat- ∗ MgSO4 7H2O 560 terns were identified and numbered. Similar restriction pat- ∗ CaCl2 2H2O 460 terns were assigned to the same group. K2SO4 150 Representative clones from each RFLP group were se- CaCO3 20 quenced (at the Institute of Biotechnology in Helsinki, Fin- Macronutrients land, or at MacroGen in Seoul, South Korea). The original NH4Cl 16.13 clones were classified on the basis of the sequencing results K2HPO4 1.55 by means of Ribosomal Database Project (RDP) taxonomic Micronutrients ∗ hierarchy classifiers (with over 97% similarities to known se- Na2WO4 2H2O0.165 ∗ quences) (Wang et al. 2007). The sequences were compared (NH4)6Mo7O24 2H2O0.44 to sequences in the GenBank, and the closest matches were KBr 0.6 selected. Representatives of the classified Proteobacteria se- KI 0.415 ∗ ZnSO4 7H2O1.435 quences were used to construct a phylogenetic tree in RDP ∗ Cd(NO3)2 4H2O0.77 Treebuilder (Cole et al. 2009). The sequences determined were ∗ Co(NO3)2 6H2O0.73 submitted to the European Nucleotide Archive and assigned ∗ CuSO4 5H2O0.625 accession numbers from HE585288 to HE585486. ∗ NiSO4(NH4)2SO4 6H2O0.99 ∗ Cr(NO3)3 9H2O0.205 ∗ KAl(SO4)2 12H2O4.74 Denaturing Gradient Gel Electrophoresis (DGGE) V2O5 0.0445 H BO 15.5 The gene encoding the bacterial 16S rRNA was amplified with 3 3 general primers specific for bacteria with a GC-clamp. The Known concentrations of Fe and Mn compounds were added separately to primers used were PRBA338f-GC: 5 CGC CCG CCG CGC the enrichment cultures. 266 Yli-Hemminki et al.

−1 with 650 mL of medium amended with 10 mg L MnSO4 (5 mM). The experimental design had six treatments, each (59 μM, 3.3 mg of dissolved Mn2+), 200 mg L−1 of Na- of which had three replicates: with living and autoclaved in- −1 acetate (CH3COONa), 80 mg L of protease peptone, 20 mg oculum as well as without inoculum, each with and without −1 L of KH2PO4, and 4 mM HEPES (4-(2-hydroxyethyl)-1- added carbon source. The inoculum was prepared from 1 g of piperazineethanesulfonicacid) (Miyata et al. 2007). The pH fresh spherical concretions, and autoclaved killed control was was adjusted to 7 with HCl and NaOH. crushed and mixed with 1 mL of artificial brackish sea water. Concretions were homogenized in the same way as for epi- Then 10 μL of the suspension was inoculated just above the fluorescence microscopy but without formaldehyde. From the FeS and Mn layers, and the pipette tip was withdrawn as the homogenized samples, 10 mL of liquid phase was added to content was expelled. Each tube was capped and incubated at ◦ three bottles while three other bottles remained as sterile con- 20 C in the dark. trols. The liquids were aerated by shaking at 140 rpm at 20◦C Subsamples from the oxidation cultures were transferred to in darkness. The flasks were compared visually and precipi- new media at 2–4-week intervals to maintain viability. To ob- tated Mn detected as dark brown precipitate. The oxidation tain larger volumes of cell material, the procedure was scaled rate was followed as decline of Mn2+ in the solution. Samples up by means of glass tubes (2.5 × 9.5 cm) with gas-permeable were taken from the bottles and filtered (0.45-μm polycar- stoppers. In larger tubes, the volume of the FeS or Mn plug bonate filter) to analyze dissolved Mn in the liquids by induc- was 5 mL and that of the upper layer was 20 mL. Cultures tively coupled plasma optical emission spectrometer (i.e., ICP- were observed via phase contrast microscopy. After the initial OES). For DAPI staining, a 5-mL sample was taken. DNA enrichments, the cells were harvested and DNA was extracted was extracted from 10 mL of filtered (0.2-μm membrane filter; from one set of replicas for cloning and sequencing. Schleicher and Schuell, Germany) sample, which was used for RFLP and the 16S rRNA gene clone library. Results Enrichment of Fe- and Mn-Oxidizing Bacteria in O2-Gradient Tubes Characterization of the Bacterial Community Structure in the Concretions A gradient method for enrichment of neutrophilic Fe- oxidizing bacteria was used, modifying the gel-stabilized (agar Spherical concretions had a pitted surface, and when they noble) gradient tubes technique described by Emerson and were cut in half, a zoned structure of alternating brown Fe Moyer (1997) and Emerson and Floyd (2005). The FeS stock and black Mn layers could be visually observed. The concre- solution was prepared after a recipe by Hanert (1992). Ac- tions studied were fragile, porous and of high water content: etate from a sterilized stock solution was added (final con- 50–55% dry weight (DW). The concretions hosted morpholog- centration: 1 mM) to the bottom-layer mixture to serve as a ically different bacteria: rods, cocci, spirilla and filamentous carbon source (Emerson and Moyer 1997). One mL L−1 of microbes with an overall average density of 6.7 × 107 cells filter-sterilized vitamin solution (Plugge 2005) and 5 mL L−1 g−1 DW. The DNA yield was higher from the outer surface of NaHCO3 solution (final concentration of 5 mM, from filter than from the interior of concretion; the two were 33 ± 13 ng sterilized 1-M stock solution) were added to the upper layer μL−1 and 11 ± 4ngμL−1, respectively. The bacterial com- medium. munity of the interior and the surface of the concretions as The same method was used for enrichment of Mn-oxidizing compared by DGGE indicated that there were several distinct ∗ bacteria, but the FeS plug was replaced with MnSO4 H2O phylotypes present (see Figure 1, pane A). Cluster analysis of

Fig. 1. (A) A DGGE proﬁle comparing the bacterial community structure in the interior (lanes 1–3) and on the surface (lanes 4–6) of Fe-Mn- concretions (here, three replicates of each). (B) Genetic similarity estimated via cluster analysis using unweighted, average clustering of the DGGE patterns shown in (panel A). The scale bar indicates the mean relative genetic similarity between the groups. Bacterial Community in Marine Fe–Mn Concretions 267

DGGE bands showed that the communities of the surface and seafloor lavas (Santelli et al. 2008), metal-rich environments the interior formed two separate clusters, in line with species such as gold-mine tailings (Nemergut et al. 2004), and the sequence differences (see Figure 1, pane B). bottom sediment of the Baltic Sea (Edlund et al. 2008). To reveal the diversity among cloned sequences, RFLP analysis was performed before the actual sequencing. At the beginning of the analysis, 169 clones in all were compared on Enrichment and Identification of Bacterial Community the basis of the differences between the restriction patterns, Members in Mn-Oxidizing Cultures and 99 clones showed unique restriction patterns indicating Inoculation with the microbial fraction of concretions en- possibly different bacteria lineages. Seventy clones could be hanced the oxidation of Mn within a week (see Figure 3), grouped into 25 RFLP groups. Several representatives of each yielding a dark precipitate, whereas no precipitation was de- RFLP-group and all of the unique clones were sequenced, tected without inoculum. One of the replicate enrichments and 55 nonsuccessful sequences were discarded. As a result, oxidized Mn twice as much as the other two, observed as in- 114 sequences were used for further identification of bacterial creasing standard error (SE) over time. Microbial growth in community members in the concretions. the gradient tubes was observed as accumulation of a dark precipitate around the inoculum after one month (see Figure 4, pane A), and the formation became faster after every transfer Identification of Bacterial Community Members to fresh culture tubes. No similar precipitation was detected in in Concretions the gradient tubes with autoclaved inocula or without inocula. The bacterial community in the spherical concretions was very Bacterial cells in the semisolid Mn-gradient media appeared diverse. Half of the cloned sequences were affiliated with Pro- club-formed (see Figure 4, pane B), and some of them were teobacteria, while nearly a third of them belonged to unknown rod-shaped and motile. In comparison, bacterial cells in the taxa (less than 90% similarity to known bacterial taxa) and liquid Mn-oxidation experiment appeared uniform and were 20% were distributed among 11 phyla (see Table 3). The closest clustered around granule-like structures, probably Mn oxides matches were uncultured bacteria originating from sea sedi- (see Figure 5). ments, Fe-rich environments, and ocean crust (see Table 3). Eighty-one cloned sequences obtained from the liquid The unclassified sequences had high scores for similarity to culture experiment were arranged into 14 RFLP groups, with uncultured bacteria found from sediments and soils (some on the maximum number of clones in any one RFLP group being hydrocarbon-polluted sites) as well as from a uranium mine 26. Clearly, some bacterial groups were enriched in this exper- (an unpublished database sequence match). iment. There were 15 clones sequenced, and after discarding All of the common Proteobacteria classes were represented of unclear sequencing results, 73 clones from the experiment in the clone library (see Figure 2): alpha (11%), gamma (11%), were classified (over 97% similarities to known sequences) delta (10%), beta (9%) and epsilon (1%). Approximately 5% of (Wang et al. 2007). The closest matches for the sequences were the clones were unclassified Proteobacteria and were excluded to cultured strains of the genera Pseudomonas, Bacillus,and from the phylogenetic tree. The closest matches to Proteobac- Sphingomonas- Sphingobium (Sphingomonads) (see Table 4) teria classes found in the concretions were obtained from, for or to uncultured bacterial clones obtained from rhizosphere example, oil-contaminated sediments (Winderl et al. 2008), and other soil. In the semisolid Mn-containing gradient,

Table 3. Bacterial 16S rRNA gene sequence type distribution, with examples, in the clone library of Fe-Mn- concretions from the GulfofFinland

Closest match Environment of Percentage Phylogenetic Sequence accession no. the closest match (n = 114) afﬁliation similarity % examples examples Reference

46.5 Proteobacteria 95–100 (Fig. 2) 3.5 Nitrospira 96–100 AF293010 Green Bay Fe-Mn micronodules Stein et al. (2001) 2.5 Acidobacteria 99–100 Gp17, GQ339135 Fe2+-rich seep Bruun et al. (2010) Gp22 2.5 Deferribacteres 99 Caldithrix AF286031 contaminated salt marsh sediment unpublished 2 Actinobacteria 97–100 EU491389 sea floor lavas from Havai’i Santelli et al. (2008) 2 Bacteroidetes 100 GQ261788 deep-sea sediment Goffredi and Orphan (2010) 2 Spirochaetes 92–97 AY605144 microbial mat unpublished 2 Chloroflexi 100 FJ746169 ocean sediment unpublished 1 Cyanobacteria 100 chloroplast FM242282 sediment unpublished 1 Firmicutes 93 EF203206 Lake Kastoria, Greece unpublished 1 OD1 93 EU283580 Lake Anderson, USA unpublished 1 WS3 100 AY114311 anoxic marine sediment unpublished 34 unclassified bacteria < 88 268 Yli-Hemminki et al.

Fig. 2. Phylogenetic tree based on 16S rRNA gene sequences of the classiﬁed Proteobacteria obtained from the clone libraries of Fe- Mn- concretions of the Gulf of Finland, Baltic Sea (‘CON’) and their most similar clone sequences or type strains (‘T’) with accession numbers. The tree was constructed by means of Ribosomal Database Project Treebuilder (weighted neighbor-joining method). The numbers at the branches are bootstrap values for 100 replications. The scale bar indicates the estimated number of base changes per nucleotide sequence position. more different bacteria, both cultured and uncultured, were comicrobia) overtook these in the absence of a carbon source. enriched, even though Sphingomonads and the class Bacilli Rheinheimera baltica was enriched without added carbon. The were not enriched in the gradient cultures. The bacteria were closest matches of the Pseudomonas sequences were not to the grouped into six classes (see Table 4). same bacterial strains as in the case of liquid Mn-oxidizing Cultured species of Pseudomonas dominated when the culture (see Figure 6). These closest matches were to cultured medium contained carbon source, but Prosthecobacter (Verru- strains originating from sediments. The addition of carbon Table 4. 16S rRNA gene sequence type distribution of bacteria from Mn- and Fe-oxidizing enrichment cultures originated from concretions of the Gulf of Finland

Enrichment conditions and Sequence Closest match Environment of the closest percentage (n of clones) Phylogenetic afﬁliation similarity % accession no. match Reference

Mn-oxidation experiment in liquid (n = 73 clones) 48 Pseudomonas sp. Gamma 99–100 EU372572 Rhizosphere soil Ofek et al. (2009) 34 Bacillus Firmicutes 100 AY072259 Soil unpublished 14 Sphingomonadaceae Alpha 100 AJ863323 Rhizosphere soil Graff and Conrad (2005) 4 unclassified Proteob. 98 DQ264498 Bioremediation of groundwater DeSantis et al. (2007) Mn-oxidation experiment in gradient with added carbon source (n = 24) 42 Pseudomonas sp. Gamma 100 EU791281 Evaporating lagoon Escalante et al. (2009) 25 Prosthecobacter sp. Verrucomicrobia 98 NR026024 Hedlund et al. (1996) 8 Devosia subaequoris Alpha 99 AM293857 Beach sediment Lee (2007) 8 Comamonadaceae Beta 99 HQ203838 Sea water unpublished 4 unclassified Alpha 99 HQ216272 Marine sewage microcosm unpublished 4 Arcobacter sp. Epsilon 93 GU088516 River Sumauma unpublished 4 Cytophagaceae Bacteroidetes 99 DQ270279 Baltic Sea proper Rieman et al. (2008) 4 Pedobacter lentus Bacteroidetes 98 AJ786785 Nitrifying inoculum unpublished Mn-oxidation experiment in gradient without added carbon (n = 19) 42 Prosthecobacter sp. Verrucomicrobia 99 AJ290012 Lake Fuchskuhle Glockner et al. (2000) 37 Pseudomonas sp. Gamma 99 FN377712 Clay sediment Srinivas et al. (2009) 5 Rheinheimera sp. Gamma 100 AJ002006 Baltic Sea Brettar et al. (2002) 5 Comamonadaceae Beta 99 EU045161 Arctic microbial mat Bottos et al. (2008) 5 Arcobacter sp. Epsilon 96 FJ469403 Oil processing seawater pipeline, Duncan et al. (2009) 5 unclassified bacteria USA Fe-oxidation experiment with added carbon source (n = 24) 42 Shewanella baltica Gamma 95–100 AJ000214 Ziemke et al. (1998) 29 Thalassolituus oleivorans Gamma 100 AJ431699 Yakimov et al. (2004) 13 Pseudomonas sp. Gamma 100 AM158413 Rhizosphere in wetland unpublished 13 Comamonadaceae Beta 100 EU356581 Agricultural soil Morales et al (2009) 4 Flavobacterium sp. Bacteroidetes 100 EU353084 Agricultural soil Morales et al. (2009) Fe-oxidation experiment without added carbon (n = 21) 29 Shewanella baltica Gamma 100 AJ000214 Ziemke et al. (1998) 29 Comamonadaceae Beta 99–100 EU356582 Agricultural soil Morales et al. (2009) 19 Pseudomonas sp. Gamma 100 AM913906 Baltic Sea Wiese et al. (2009) 10 unclassified Beta 98 AB241013 River sediment unpublished 10 Flavobacterium sp. Bacteroidetes 100 HQ203838 Sea water unpublished 269 5 unclassified Alpha 95 DQ337073 Subsurface water, Kalahari unpublished 270 Yli-Hemminki et al.

Fig. 3. The concentrations of dissolved Mn2+ in liquid culture media with and without inocula from the Fe-Mn-concretions. The error bars are standard errors (SE) calculated from three replicates of each treatment. to Mn-gradient media caused differences again in the closest Fig. 5. Micrograph of a DAPI-stained sample from Mn oxidation matches for Pseudomonas sequences between the experiments. in a liquid culture experiment with microbes extracted from concretions from the Baltic Sea. Enriched cells were clustered around granule-like structures that probably consisted of Mn oxides. Enrichment and Identification of Bacterial Community Members in Fe-Oxidizing Cultures Microbial growth in FeS-containing gradient tubes was observed as formation of a reddish-brown layer around the location of the original inoculum within a week from inoculation (see Figure 7, pane A). Addition of carbon did not significantly affect the visible formation of brown layers. It is noteworthy The Fe gradient favored the enrichment of Shewanella that bands of Fe oxides did not form with autoclaved inocula baltica, and with added carbon source the Thalassolituus or without inocula but Fe oxides slowly diffused throughout oleivorans too was enriched. Pseudomonas sp., uncultured Co- the upper layer. The growth layer formed more quickly, in 3 mamonadaceae and Flavobacterium sp. were enriched with or to 4 days, after every transfer to fresh culture tubes. Bacterial without added carbon source. Pseudomonas sequences were cells were commonly filamentous and attached to Fe-oxide not as dominant as in Mn-containing media, and the closest particles (the image in pane B of Figure 7 was obtained after matches in the databank of sequences were found from the the third transfer to fresh culture tubes). Baltic Sea and agricultural soil (see Figure 6).

Fig. 4. (A) Gradient growth tubes with MnSO4-agarose plug on the bottom with and without carbon source (tubes 1 and 2). Dark precipitate forming around bacteria inoculated from the Baltic Sea concretions indicated the formation of Mn oxide. Tubes 3 and 4 were autoclaved inocula with and without carbon source. (B) Club-formed bacteria growing in opposing gradients of oxygen and MnSO4 with carbon source as viewed via phase contrast microscopy. The scale bar is 10 μm. Bacterial Community in Marine Fe–Mn Concretions 271

Fig. 6. Phylogenetic tree based on 16S rRNA gene sequences of Pseudomonas obtained from the clone libraries of Fe- and Mn- oxidizing enrichment cultures originated from concretions of the Gulf of Finland, Baltic Sea, and their most similar clone sequences or type strains (‘T’) with accession numbers. The letter ‘C’ in the clone names indicates that there was an added carbon source in the enrichment. The tree was constructed by means of Ribosomal Database Project Treebuilder (weighted neighbor-joining method). The numbers at the branches are bootstrap values for 100 replications. The scale bar indicates the estimated number of base changes per nucleotide sequence position.

Discussion that the surface layer had a greater diversity, whereas some strains of bacteria were more pronounced in the interior of Fragility and the high water content (about 50% of DW) in- the concretion. The smaller DNA yield from the interior may dicate that the spherical concretions from the Gulf of Fin- indicate less abundant bacteria in the interior than on the sur- land have a porous structure. A similar pitted surface and face of the concretion even if DNA extraction is not a reliable alternating, concentric ring structure have been observed in method for determining bacterial abundance. The enrichment soil Fe-Mn nodules (Gasparotos et al. 2005). These structural of taxa could result from chemical gradients at the interface characteristics may go some way toward explaining why the between the concretion surface and interior, in the variation concretions in the Gulf of Finland were colonized by more of microenvironments provided by the concretions. than 107 microbial cells g−1 DW and the bacterial diversity The closest matches to the cloned 16S rRNA gene se- was high. Our DNA extractions and DGGE results indicated quences of the concretion bacteria were to sequences obtained that bacteria colonize both the surface and the porous inte- from similar types of environments. With regard to the physio- rior of an individual concretion. The DGGE proﬁles showed logical properties of the bacteria found, it is worthy of note that 272 Yli-Hemminki et al.

Fig. 7. (A) Bacteria from Fe-Mn-concretions grew as layers in opposing gradients of FeS (dark bottom plug) and oxygen with and without added Na-acetate (1 mM) (tubes 1 and 2 from the left, respectively). Tubes 3 and 4 were uninoculated controls with and without Na-acetate. Tubes 5 and 6 were autoclaved inoculates with and without Na-acetate. (B) Bacteria viewed by means of phase contrast microscopy after the third transfer to fresh culture tubes. Filamentous bacteria grew attached to Fe oxide particles in opposing gradients of oxygen and FeS with carbon source. The bright spots in the picture appeared reddish brown in actual view. Thescalebaris10μm. many of the known Mn- and Fe-oxidizers belong to the alpha-, with members of Sphingomonadaceae known to biodegrade beta- and gammaproteobacteria whereas the Fe-reducers, such contaminants (Zheng et al. 2011). In the semisolid Mn media, as Geobacter, belong to the deltaproteobacteria (Weber et al. ubiquitous and versatile Pseudomonas still dominated, and it 2006). Half of the cloned sequences from the concretions be- succeeded in the Fe gradient as well. longed to Proteobacteria and represented relatively evenly the Different enrichment conditions resulted in differences be- most common classes. Yet, 5% of the cloned sequences in our tween the closest matches to Pseudomonas sequences. Slowly study were unclassified Proteobacteria and concretions were growing appendage-forming Prosthecobacter was enriched not dominated by the known Fe- and Mn-oxidizing or Fe- and only with Mn. Prosthecobacter belongs to the phylum Ver- Mn-reducing bacteria. rucomicrobia, which is one of the most abundant taxa of the As in our study, the bacteria strains directly related to the bacterial community in the water column of the Baltic Sea cycling of metals have been found to be less abundant in fresh- (Herlemann et al. 2011). Prosthecobacter strains form buds water Fe-Mn micronodules of the Green Bay (Stein et al. (prosthecae) on filaments and are often attached to each other 2001). The proportion of unclassified bacteria was consider- after cell division (Hedlund et al. 1997). There are no previous able from the concretions of the Baltic Sea, and their potential reports on their ability to oxidize Mn. Hyphal bacteria have abilities to oxidize Fe and Mn remained unknown. However, been seen in the enrichment cultures obtained from the surface sometimes the most dominant bacteria are not necessarily layers of the concretions from the western Baltic Sea (Ghiorse metabolically active (Edlund et al. 2008). A relatively small and Hirsch 1982). proportion of the bacterial community might be responsible In our experiment, Prosthecobacter strains were enriched for the reaction actually observed. The high bacterial diversity even without added carbon, although they have failed to grow in the concretions was in contrast with the assumption of con- on one-carbon or two-carbon compounds in earlier studies cretions being selective and unfavorable habitats. A greater (Hedlund et al. 1997). It may be that, decaying bacterial number of cloned sequences in the analysis would probably cells provided carbon source for their growth. In our experi- have increased the frequency of smaller bacteria groups, which ment, the formation of black Mn-oxides was accelerated after may now be underestimated or not detected, as could be the each transfer of bacteria to a fresh gradient tube especially case even if the primers used had a broad coverage. when they had Na-acetate as a carbon source. Several Mn2+- Our oxidation experiments demonstrated that an individual oxidizing heterotrophic strains, mostly alpha- and gamma- concretion serves as habitat for a diverse bacterial community Proteobacteria, have also been cultured from Fe-Mn crusts of having potential for the growth of Mn- and Fe-oxidizers. The Lo‘ihi¯ Seamount, in Hawaii (Templeton et al. 2005). cultivation media with Mn or Fe, agitation, and addition of The new candidatus class belonging to the neutrophilic carbon resulted in differences in enriched bacteria species. The Fe-oxidizing bacteria ‘ζ -Proteobacteria’ (zeta) (Emerson and formation of oxidized Fe and Mn was observed around living Moyer 2002; Emerson et al. 2007) were not found from the inocula either with or without the added carbon source. The concretions from the Baltic Sea. Despite the growth of or- main selective factor was the added reduced metal. Mainly ange precipitate indicating the formation of Fe oxides, the rapidly growing known Mn-oxidizers such as Pseudomonas bacteria strains found did not fall into categories of known Fe- and Bacillus were enriched in the liquid culture media, along oxidizers. On the contrary, the Shewanella species enriched are Bacterial Community in Marine Fe–Mn Concretions 273 considered to be heterotrophic Fe-reducers. Here Shewanella Brettar I, Christen R, Hofle MG. 2002. Rheinheimera baltica gen. nov., grew also without added carbon. Added carbon enhanced sp. nov., a blue-coloured bacterium isolated from the central Baltic the heterotrophic growth of aerobic Thalassolituus oleivorans, Sea. Int J Syst Evol Microbiol 52:1851–1857. known to utilize petroleum hydrocarbons (Yakimov et al. Bruun AM, Finster K, Gunnlaugsson HP, Nornberg P, Friedrich MW. 2010. A Comprehensive investigation on iron cycling in a freshwater 2004). In these experiments, we investigated the immediate seep including microscopy, cultivation and molecular community 2+ enrichment of Fe-oxidizing bacteria in the oxygen-Fe gradi- analysis. Geomicrobiol J 27:15–34. ent after the first inoculation with concretion suspension. It is Cole JR, Wang Q, Cardenas E et al. 2009. The Ribosomal Database possible that the aerobic species grew on the top layer of the Project: improved alignments and new tools for rRNA analysis. gradient tube whereas Shewanella, capable of accumulating Nucl Acids Res 37:D141–145. and reducing Fe, could grow in the anoxic layer. Shewanella is DeSantis TZ, Brodie EL, Moberg JP, Zubieta IX, Piceno YM, Andersen known to exhibit enrichment in microaerophilic conditions. GL. 2007. High-density universal 16S rRNA microarray analysis reveals broader diversity than typical clone library when sampling the environment. Microb Ecol 53:371–383. Duncan KE, Gieg LM, Parisi VA, Tanner RS, Tringe SG, Bristow J, Suflita JM. 2009. Biocorrosive thermophilic microbial commu- Conclusions nities in Alaskan North Slope oil facilities. Environ Sci Technol 43:7977–7984. Each individual spherical concretion forms a microcosm for Edlund A, Hardeman F, Jansson JK, Sjoling S. 2008. Active bacterial microbial life colonized by diverse and partly unknown bac- community structure along vertical redox gradients in Baltic Sea teria. Our data suggested that not only is the surface of the sediment. Environ Microbiol 10:2051–2063. concretion colonized by bacteria but they are also found from Edwards KJ, Bach W, Rogers DR. 2003a. Geomicrobiology of the the interior of the concretion. The diverse bacterial commu- ocean crust: a role for chemoautotrophic Fe-bacteria. Biol Bull 204:180–185. nity indicated that concretions may form a specific location Edwards KJ, Rogers DR, Wirsen CO, McCollom TM. 2003b. for a complicated network of bacterial metabolism. The com- Isolation and characterization of novel psychrophilic, neu- munity resembles that of the ocean crust and other Fe-Mn- trophilic, Fe-oxidizing, chemolithoautotrophic alpha- and gamma- rich environments. The bacteria hosted by concretions oxi- proteobacteria from the deep sea. Appl Environ Microbiol dized Mn and affected the formation of Fe oxides, which is in 69:2906–2913. concert with the view that the formation of concretions is par- Ehrlich HL. 1963. Bacteriology of manganese nodules. I. Bacterial action on manganese in nodule enrichments. Appl Microbiol 11:15– tially biogenous. It is noteworthy that the known Mn- and Fe- 19. oxidizing bacteria did not dominate the bacterial community Emerson D, Floyd MM. 2005. Enrichment and isolation of iron-oxidizing of concretions. The abundance of concretions and their phys- bacteria at neutral pH. Meth Enzymol 397:112–123. ical and geochemical properties, together with the diversity Emerson D, Moyer C. 1997. Isolation and characterization of novel iron- seen and the ability of bacteria to participate in the oxidation oxidizing bacteria that grow at circumneutral pH. Appl Environ of Mn and Fe, indicate that the concretions and the microbes Microbiol 63:4784–4792. they host may significantly affect the biogeochemical cycling Emerson D, Moyer CL. 2002. Neutrophilic Fe-oxidizing bacteria are abundant at the Loihi seamount hydrothermal vents and play of elements at the seafloor of the Baltic Sea. a major role in Fe oxide deposition. Appl Environ Microbiol 68:3085–3093. Emerson D, Revsbech NP. 1994. Investigation of an iron-oxidizing microbial mat community located near Aarhus, Denmark: field studies. Acknowledgments Appl Environ Microbiol 60:4022–4031. Emerson D, Rentz JA, Lilburn TG, Davis RE, Aldrich H, Chan C, Moyer This work was financially supported by the Maj and Tor CL. 2007. A novel lineage of Proteobacteria involved in formation Nessling Foundation and the K.H. Renlund Foundation. We of marine Fe-oxidizing microbial mat communities. PLOS One 2(8): also thank Aarno Kotilainen from the Geological Survey of e667. Finland and Anna-Leena Downie of the VELMU project, Escalante AE, Caballero-Mellado J, Martinez-Aguilar L, Rodriguez- Verdugo A, Gonzalez-Gonzalez A, Toribio-Jimenez J, Souza V. from the Finnish Environment Institute, along with the crew 2009. Pseudomonas cuatrocienegasensis sp. nov., isolated from an of R/V Muikku and R/V Aranda, for supplying concretions. evaporating lagoon in the Cuatro Cienegas valley in Coahuila, Mex- The authors wish to extend gratitude to Nina Hurme, Ludi- ico. Int J Syst Evol Microbiol 59:1416–1420. vine Fleury and Sanja Karlsson for their contribution to the Gasparatos D, Tarenidis D, Haidouti C, Oikonomou G. 2005. Micro- experimental and DNA work. scopic structure of soil Fe-Mn nodules: environmental implications. Environ Chem Lett 2:175–178. Ghiorse WC, Hirsch P. 1982. Isolation and properties of ferromanganese- References depositing budding bacteria from Baltic Sea ferromanganese concretions. Appl Environ Microbiol 43:1464–1472. Anufriev GS, Blinov LN, Boltenkov BS, Arif M. 2005. Chemical and Glasby GP, Stoffers P, Sioulas A, Thjissen T, Friedrich G. 1982. Man- isotope composition of Baltic iron-manganese concretions. Tech ganese nodule formation in the Pacific Ocean: a general theory. Phys 50:663–665. Geo-Mar Lett 2:47–53. Baturin GN, Dubinchuk VT. 2009. Composition of ferromanganese nod- Glockner FO, Zaichikov E, Belkova N, Denissova L, Pernthaler J, Pern- ules from Riga Bay (Baltic Sea). Oceanology 49:111–120. thaler A, Amann R. 2000. Comparative 16S rRNA analysis of lake Bottos EM, Vincent WF, Greer CW, Whyte LG. 2008. Prokaryotic di- bacterioplankton reveals globally distributed phylogenetic clusters versity of arctic ice shelf microbial mats. Environ Microbiol 10:950– including an abundant group of actinobacteria. Appl Environ Mi- 966. crobiol 66:5053–5065. 274 Yli-Hemminki et al.

Goffredi SK, Orphan VJ. 2010. Bacterial community shifts in taxa and Rieman L, Leitet C, Pommier T, Simu K, Holmfeldt K, Hagstrom¨ A. diversity in response to localized organic loading in the deep sea. 2008. The native bacterioplankton community in the central Baltic Environ Microbiol 12:344–363. Sea is influenced by freshwater bacterial species. Appl Environ Mi- Graff A, Conrad R. 2005. Impact of flooding on soil bacterial commu- crobiol 74:503–515. nities associated with poplar (Populus sp.) trees. FEMS Microbiol Santelli CM, Orcutt BN, Banning E, Bach W, Moyer CL, Sogin ML, Ecol 53:401–415. Staudigel H, Edwards KJ. 2008. Abundance and diversity of micro- Hanert HH. 1992. The genus Gallionella. In: Balows A, Truper¨ HG, bial life in ocean crust. Nature 453:653–656. Dworkin M, Harder W, Schleifer K-H., eds. The Prokaryotes, a Srinivas TN, Nageswara Rao SS, Vishnu Vardhan Reddy P, Pratibha MS, Handbook on the Biology of Bacteria: Ecophysiology, Isolation, Sailaja B, Kavya B, Hara Kishore K, Begum Z, Singh SM, Shivaji S. Identification, Applications, 2nd ed. Vol. 4. Springer-Verlag: New 2009. Bacterial diversity and bioprospecting for cold-active lipases, York, p4082–4088. amylases and proteases, from culturable bacteria of Kongsfjorden Hauck S, Benz M, Brune A, Schink B. 2001. Ferrous iron oxidation by and Ny-Alesund, Svalbard, Arctic. Curr Microbiol 59:537–547. denitrifying bacteria in profundal sediments of a deep lake (Lake Stein LY, La Duc MT, Grundl TJ, Nealson KH. 2001. Bacterial and Constance). FEMS Microbiol Ecol 37:127–134. archaeal populations associated with freshwater ferromanganous He J, Zhang L, Jin S, Zhu Y. 2008. Bacterial communities inside and sur- micronodules and sediments. Environ Microbiol 3:10–18. rounding soil iron-manganese nodules. Geomicrobiol J 25:14–24. Straub KL, Benz M, Schink B, Widdel F. 1996. Anaerobic, nitrate- Hedlund BP, Gosink JJ, Staley JT. 1997. Verrucomicrobia div. nov., a dependent microbial oxidation of ferrous iron. Appl Environ Mi- new division of the Bacteria containing three new species of Pros- crobiol 62:1458–1460. thecobacter. Anton von Leeuwen 72:29–38. Tarkiainen E, Rinne I. 1973. Algal assay procedure used in Water Con- Herlemann DPR, Labrenz M, Jurgens¨ K, Bertilsson S, Waniek JJ, servation Laboratory. Helsinki: City of Helsinki Building Office, Andersson AF. 2011. Transitions in bacterial communities along Street Construction Department. Reports of the Water Conserva- the 2000 km salinity gradient of the Baltic Sea. ISME J 5:1571– tion Laboratory 12: 6. 1579. Tebo BM, Bargar JR, Clement BG, Dick GJ, Murray KJ, Parker D, Ver- Ingri J. 1985. Geochemistry of ferromanganese concretions and associ- ity R, Webb SM. 2004. Biogenic manganese oxides: Properties and ated sediments in the Gulf of Bothnia. PhD thesis, Lulea˚ University mechanisms of formation. Annu Rev Earth Planet Sci 32:287–328. of Technology, Sweden. ISSN 0348-8373. Tebo BM, Emerson S. 1986. Microbial manganese(II) oxidation in Konhauser KO, Kappler A, Roden EE. 2011. Iron in microbial the marine environment: A quantitative study. Biogeochemistry metabolism. Elements 7:89–93. 2:149–161. Kuosmanen T, Peltola M, Raulio M, Pulliainen M, Laurila T, Selin Tebo BM, Villalobos M. 2005. Introduction: Advances in the geomi- J-F, Huopalainen H, Salkinoja-Salonen M. 2006. Effect of polar- crobiology and biogeochemistry of manganese and iron oxidation. ization on manganese biofouling of heat exchanger surfaces. In: Geomicrobiol J 22:77–78. Muller-Steinhagen¨ H, Malayeri MR, Watkinson AP, editors. ECI Templeton AS, Staudigel H, Tebo BM. 2005. Diverse Mn(II)-oxidizing Symposium Series, vol RP2: Proceedings of 6th International Con- bacteria isolated from submarine basalts at Loihi seamount. Ge- ference on Heat Exchanger Fouling and Cleaning – Challenges and omicrobiol J 22:127–139. Opportunities; Kloster Irsee, Germany, pp. 283–288. Teske A, Wawer C, Muyzer G, Ramsing NB. 1996. Distribution of Lee SD. 2007. Devosia subaequoris sp. nov., isolated from beach sediment. sulfate-reducing bacteria in a stratified fjord (Mariager Fjord, Den- Int J Syst Evol Microbiol 57:2212–2215. mark) as evaluated by most-probable-number counts and denatur- Marcus MM, Manceau A, Kersten M. 2004. Mn, Fe, Zn and As spe- ing gradient gel electrophoresis of PCR-amplified ribosomal DNA ciation in a fast-growing ferromanganese marine nodule. Geochim fragments. Appl Environ Microbiol 62:1405–1415. Cosmochim Acta 68:3125–3136. Thamdrup B, Glud RN, Hansen JW. 1994. Manganese oxidation and Mason OU, Di Meo-Savoie CA, Van Nostrand JD, Zhou J, Fisk MR, in situ manganese fluxes from a coastal sediment. Geochim Cos- Giovannoni SJ. 2009. Prokaryotic diversity, distribution and in- mochim Acta 58:2563–2570. sights into their role in biogeochemical cycling in marine basalts. Wang Q, Garrity GM, Tiedje JM, Cole JR. 2007. Na¨ıve Bayesian clas- ISME J 3:231–242. sifier for rapid assignment of rRNA sequences into new bacterial Miyata N, Sugiyama D, Tani Y, Tsuno H, Seyama H, Sakata M, taxonomy. Appl Environ Microbiol 73:5261–5267. Iwahori K. 2007. Production of biogenic manganese oxides by Wang X, Schroder¨ HC, Wiens M, Shloßmacher U, Muller¨ WEG. repeated-batch cultures of laboratory microcosms. J Biosci Bioeng 2009. Manganese/polymetallic nodules: micro-structural character- 105:432–439. ization of exolithobiontic- and endolithobiontic microbial biofilms Morales SE, Cosart TF, Johnson JV, Holben WE. 2009. Extensive phy- by scanning electron microscopy. Micron 40:350–358. logenetic analysis of a soil bacterial community illustrates extreme Weber K, Achenbach LA, Coates JD. 2006. Microorganisms pumping taxon evenness and the effects of amplicon length, degree of cover- iron: anaerobic microbial iron oxidation and reduction. Nature Rev age, and DNA fractionation on classification and ecological param- Microbiol 4:752–764. eters. Appl Environ Microbiol 75:668–675. Wiese J, Thiel V, Nagel K, Staufenberger T, Imhoff JF. 2009. Diver- Nemergut DR, Martin AP, Schmidt SK. 2004. Integron diversity of antibiotic-active bacteria associated with the brown alga sity in heavy-metal-contaminated mine tailings and inferences Laminaria saccharina from the Baltic Sea. Mar Biotechnol 11:287– about integron evolution. Appl Environ Microbiol 70:1160– 300. 1168. Winderl C, Anneser B, Griebler C, Meckenstock RU, Lueders T. 2008. Ofek M, Hadar Y, Minz D. 2009. Comparison of effects of compost Depth-resolved quantification of anaerobic toluene degraders and amendment and of single-strain inoculation on root bacterial com- aquifer microbial community patterns in distinct redox zones of a munities of young cucumber seedlings. Appl Environ Microbiol tar oil contaminant plume. Appl Environ Microbiol 74:792–801. 75:6441–6450. Winterhalter B. 1966. Pohjanlahden ja Suomenlahden rautamangaani- Øvreas˚ L, Forney L, Daae FL, Torsvik V. 1997. Distribution of bacterio- saostumista. Geoteknillisia¨ julkaisuja No. 69, Geologian tutkimus- plankton in meromictic Lake Saelenvannet, as determined by dena- laitos, Otaniemi, Finland. turing gradient gel electrophoresis of PCR-amplified gene fragments Yakimov MM, Giuliano L, Denaro R, Crisafi E, Chernikova TN, Abra- coding for 16S rRNA. Appl Environ Microbiol 63:3367–3373. ham W-R, Luensdorf H, Timmis KN, Golyshin PN. 2004. Thalas- Plugge CM. 2005. Anoxic media design, preparation and considerations. solituus oleivorans gen. nov., sp. nov., a novel marine bacterium that Meth Enzymol 397:3–16. obligately utilizes hydrocarbons. Int J Syst Ev Microbiol 54:141–148. Bacterial Community in Marine Fe–Mn Concretions 275

Zhamoida V, Grigoriev A, Gruzdov K, Ryabchuk D. 2007. The influ- Zhang L-M, Liu F, Tan W-F, Feng X-H, Zhu Y-G, He J. 2008. Microbial ence of ferromanganese concretions-forming processes in the east- DNA extraction and analyses of soil iron-manganese nodules. Soil ern Gulf of Finland on the marine environment. In: Vallius H, ed. Biol Biochem 40:1364–1369. Holocene sedimentary environment and sediment geochemistry of Zheng G, Selvam A, Wong JWC. 2011. Rapid degradation of lindane (γ - the Eastern Gulf of Finland, Baltic Sea. Geological Survey of Fin- hexachlorocyclohexane) at low temperature by Sphingobium strains. land, Special Paper 45:21–32. Espoo, Finland. Int Biodeter Biodegr 65:612–618. Zhang FS, Lin CY, Bian LZ, Glasby GP, Zhamoida VA. 2002. New Ziemke F, Hofle MG, Lalucat J, Rossello-Mora R. 1998. Reclas- evidence for the biogenic formation of spheroidal ferromanganese sification of Shewanella putrefaciens Owen’s genomic group II concretions from the Eastern Gulf of Finland, Baltic Sea. Baltica as Shewanella baltica sp. nov. Int J Syst Bacteriol 48:179– 15:23–29. 186.

Iron–manganese concretions contribute to benthic release of phosphorus and arsenic in anoxic conditions in the Baltic Sea

Pirjo Yli–Hemminki*1, Timo Sara–aho2, Kirsten S. Jørgensen1, Jouni Lehtoranta1

1Marine Research Centre, Finnish Environment Institute, P.O.B. 140, FI–00251 Helsinki, Finland 2Laboratory Centre, Finnish Environment Institute, P.O.B. 140, FI–00251 Helsinki, Finland

*Corresponding author Pirjo Yli–Hemminki 1Marine Research Centre, Finnish Environment Institute P.O.B. 140 FI–00251 Helsinki, Finland Tel.: +358(0)295251764; Fax: +358(0)9495913 e–mail: pirjo.yli–[email protected]

Abstract Purpose: Deposits of iron–manganese (Fe, Mn) concretions form a large storage of phosphorus (P) and arsenic (As) particularly under pressure of oscillating oxygen conditions in the eutrophic Gulf of Finland, the Baltic Sea. There is a poor understanding of how anaerobic microbial processes regulate the cycling of elements in the concretions. The objective of this study was to highlight how the microbial processes control the release of elements from the concretions to brackish water during anoxia. Materials and methods: Spherical concretions were collected from oxic bottoms from the Gulf of Finland in the summer. Concretions and autoclaved controls were incubated in anoxic artificial brackish seawater with and without labile carbon, plus supplied with ammonium at 5, 10 and 20 °C for fifteen weeks. Concentrations of Fe, Mn, P and As were measured (inductively coupled plasma optical emission spectrometer) from the intact concretions, and the release of the elements was measured during the experiment. Also the consumption of the added ammonium and organic carbon and the formation of dissolved inorganic carbon were measured. Results and discussion: At near–bottom temperature 5 °C, the concretions released at highest 0.12, 0.42, 0.02 and 0.0002 μmol g-1 d-1 of Fe, Mn, P, and As, respectively. The rates were significant only in the microcosms with added labile carbon, and only minor proportions (0.1–0.4%) of their total contents were released during the incubations. The concretions removed completely the supplied ammonium only without carbon addition. During prolonged anoxia concretion deposit may form a local hot spot for the microbial reduction of Fe and Mn and release significant amounts of P and As, and participate in N cycling besides the bottom sediments of the Gulf of Finland. However, the concretions may maintain their binding capacity for P and As longer than the fine–grained organic rich sediment during anoxia. Conclusions: Concretion deposits may form a temporal source of bioavailable P having ecological significance in the Gulf of Finland when concretions have access to labile organic carbon under anoxia. Concretions from the Baltic Sea, the Oceans, lakes and soils show similarities in their high content of redox–sensitive metals. Anaerobic microbial processes may also affect the stability of concretions from the different environments but the outcome may depend on the variability of ambient geochemical conditions.

Keywords Eutrophication • Gulf of Finland • Microbial reduction • Nodules

1 Introduction Iron–manganese rich particulate precipitates (hereafter Fe–Mn concretions) are formed in soils, lakes and brackish sedimentary systems and not least on the ocean seafloor (see Table A1, Electronic Supplementary Material). In the main basin of the Baltic Sea the occurrence of concretions is minor, but large deposits are found in the Gulf of Finland (GoF), the Gulf of Bothnia (GoB) and the Gulf of Riga (GoR). Dense deposits of spheroidal concretions may cover about 10% of the bottom area of the GoF and the GoB (Winterhalter 1966; Ingri 1985). The estimated density in the GoR is 17 kg m-2 and in the GoB 15 kg m-2 (Ingri 1985; Baturin and Dubinchuk 2009) and it can be as high as 24 kg m- 2 in the eastern GoF (Glasby et al. 1997). The Fe–Mn concretion deposits contain high concentrations of phosphorus (P) and arsenic (As) in the GoF (Ingri 1985; Baturin 2009; Baturin and Dubinchuk 2009). It has been proposed that concretions form a significant storage of P (Savchuk 2000), and it has been estimated that the concretion bottoms in the eastern GoF alone may contain about 175 000 tons of P (Vallius et al. 2011). The value is ten times more than the 17 000 ton pool of P bound to Fe in the surface sediments of the accumulation bottoms which suffer frequent anoxia (Lehtoranta and Pitkänen 2003). It is evident that concretions do not withstand long–term anoxia, because they are not found from areas of high sedimentation rates (Ingri 1985), and the geochemical signs of their decay are found when they are buried under sediment layers greater than 50–100 mm (Zhamoida et al. 2007). Anoxic events caused mainly by fluctuating halocline (Pitkänen et al. 2003; Haahti and Kangas 2006) reach occasionally the areas covered by concretion deposits in the GoF (Hansson et al. 2011; Vallius et al. 2011). We consider that in anoxia are geochemical changes developed in the concretions, and that these changes are mainly driven by microbial processes. A close relationship between microbes and concretions has already been revealed by morphologically and taxonomically diverse bacteria found from both the surface and interior of concretions (Ghiorse and Hirsch 1982; Yli–Hemminki et al. 2014). Previous studies have also indicated that microbes are likely to contribute to the formation of Fe– and Mn–oxides in the concretions (Yli–Hemminki et al. 2014), and thus they are likely to participate in the sequestration of P and As to Fe–oxides (Marcus et al. 2004; Chen et al. 2006). Binding of As to Fe3+ oxides (Farquhar et al. 2002) may reduce its bioavailability, and thus toxicity. Mn oxidation is greatly enhanced by bacteria (Tebo and Emerson 1986), observed also in the experiments with concretion bacteria (Ehrlich 1968; Yli–Hemminki et al. 2014). It is evident that Fe and Mn in the concretions are in the form of oxides, and they may serve as electron acceptors for metal–reducing bacteria when conditions turn anoxic. The aim of the present study was to highlight how anaerobic microbial processes turn Fe–Mn concretions unstable, and how the additions of labile organic carbon (C) and + ammonium (NH4 ) affect the release of Fe, Mn, P and As, and behaviour of nitrogen (N) in anoxia. The effects of temperature were studied to test the response of concretions on temperatures higher than in near–bottom water. The release rates of Fe, Mn, P and As from spherical Fe–Mn concretions incubated in anoxia are reported and compared to the published values reported earlier from sediments. Also, the contribution of microbial reduction of concretion metals to the benthic release of P and As and cycling of N in the GoF is discussed. Additionally, the elemental composition of the concretions of the GoF and the Baltic Sea are compared to the previously published values of concretions found from oceans, lakes and soils to highlight whether the concretions from different geographical areas and environments have similar potential to release sequestered elements 3 to their surroundings as in the anoxic brackish system. The ecological significance of microbe–mineral interactions coupled to the release of P and As from the concretions has not been considered previously for concretion–rich regions.

2 Material and methods 2.1. Sampling and experimental setup of incubations Concretions were collected from the bottoms of water depths which are occasionally anoxic in the GoF (three sampling sites, Table 1). The concretions were spherical and their diameter varied within the sampling site. Concretions were sampled by using a van Veen grab sampler during the Finmarinet cruise in July 2010. Remotely operated vehicle equipped with video camera (ROV), run prior to sampling and showed that the areas were covered by spherical Fe–Mn concretions. The near–bottom water of the sampling sites was oxic, and the bottom sediment was dwelled by burrowing fauna. Concretions were gently rinsed, and they were not let to dry. Concretions were stored in near–bottom water in gas– permeable plastic bags at 5 °C in darkness. The concretions from the three sampling sites were pooled and incubated in sealed flasks for 15 weeks to simulate anoxic conditions in the bottoms of the GoF. Concretions in the experiments were treated separately from their surrounding sediment in order to test the response of the concretions alone. Mixtures of wet concretions were weighted (10 ± 1 g wet weight) into sterile wide–mouthed bottles (total volume ca. 110 ml). Artificial brackish sea water (salinity 4.7‰) with 100 mL in each unit was the basis for a growth medium (described in Yli–Hemminki et al. 2014), amended with 1 mg L-1 vitamin–B solution (Plugge 2005). Each unit was flushed with N2 –gas for at least 5 minutes to purge oxygen from the solution, and the bottles were sealed with air–tight top stoppers. The medium did not contain added Fe, Mn, P or As, so the elements found in dissolved form in the studied 2+ solution originated from the concretions. The sulphate (SO4 ) concentration of the medium was 300 mg L-1. The incubations were not supposed to be N–limited. Therefore, ammonium chloride (NH4Cl) was added as N source for microbes, and its’ concentration varied between 3–4 mg L-1 after autoclaving the growth medium. The concentrations of + NH4 close to the measured values before incubations are commonly found in the porewaters of the GoF (Lehtoranta and Pitkänen 2003). Sets of concretions (5 flasks for each set) were incubated with and without the added carbon (sodium acetate trihydrate, CH3COONa*3H2O, 10 mM) at temperatures of 5, 10 and 20 °C in darkness. In the coastal GoF a temperature close to 4 °C is normal in near–bottom water and reaches the highest value of about 14°C only during the autumn turnover (Leppäranta and Myrberg 2009). The sterile control experiments (2 flasks for each set) were prepared by autoclaving concretions at 120 °C two successive periods for 30 minutes. The solutions for analyses of Fe2+, dissolved total Fe, Mn, P and As were sampled with a sterile needle–filter–syringe (pore size 0.45μm polycarbonate filter) combination. The dissolution of Fe2+ was measured from the same incubation unit to follow its release from the concretions (weeks 0, 2, 5, 6, 7, 8, 9, 10, 11, 12, 14). Destructive sampling was carried out to determine the dissolution of total Fe, Mn, P and As, consumption of + dissolved organic carbon (DOC) and NH4 , and the formation of dissolved inorganic carbon (DIC) during the incubations. One incubation unit was sacrificed at a time (weeks 0, 2, 5, 6, 15). Autoclaved incubations were sampled in the beginning and at the end of the experiment.

2.2 Geochemical analyses of solid concretions and elements released in the incubations Dry weight of concretions was determined by freeze drying (9 replicates i.e. 3 from each three sampling sites), and loss on ignition (LOI) by igniting concretions in a furnace at 550 °C for two hours. Total concentrations of Fe, Mn, P and As were analysed from the freeze– dried concretions homogenized in an agate mortar. Aliquots (0.05 g) were digested in Teflon vessels with 6 mL HNO3 and 1 mL H2O2 in a microwave oven. Samples were heated at 180 °C for 20 minutes. After cooling, samples were diluted (MQ–H2O) to 30 mL in graduated polypropylene test tubes and stored in dark at room temperature until analysis. Determinations of As, Fe, Mn and P were performed with an Inductively Coupled Plasma Optical Emission Spectrometer (ICP–OES Varian Vista, Varian Inc.). Further sample dilution was necessary for Fe and Mn due to their high concentrations. The ferrozine–method (Stookey 1970) was used to analyse the Fe2+ in solution. The concentrations of total dissolved Fe, Mn, P and As were analysed from the ambient water. The samples were diluted and acidified with 0.5 % (v/v) HNO3 before analyses. Determinations of Mn and P were performed with an ICP–OES. The measurement technique was not sensitive enough for As and some of the Fe concentrations in the incubations. Therefore, As and Fe were determined with an Inductively Coupled Plasma Mass Spectrometer (ICP–MS, PerkinElmer Elan DRC II or Thermo iCAP Q) utilizing Dynamic Reaction Cell technology (DRC, PerkinElmer) or Kinetic Energy Discrimination (KED, Thermo) to reduce polyatomic interferences. Rhodium was added as an internal standard with a final concentration of 10 μg L-1 before measurement. The DOC samples were filtered through a polycarbonate filter (pore size 0.45μm) and introduced into the combustion tube where the C content was oxidized and measured with an infrared (IR) detector as CO2. Analysis of DIC was carried out by converting inorganic C in the sample to CO2 using phosphoric acid. The CO2 was then volatilized and detected by the IR–detector. The methods are based on the SFS–EN 1484 standard for + water analysis (European standard implemented in Finland). The concentration of NH4 was analysed by measuring the absorbance of a formed coloured compound, indophenol, at + 630 nm. Our NH4 analysis is based on the standard SFS–EN ISO 11732 (international standard implemented in Finland). The amounts of released element g-1 concretion DW per unit of time are calculated on the basis of the concretions weighted and the amount of elements released to the -1 ambient solution. The release rates for the elements were calculated by: (μmol g DW2 - -1 μmol g DW1)/ (t2 - t1), where “t” is sampling time.

3 Results 3.1 Physical and geochemical properties of concretions Spherical concretions were porous according to their average dry weight (DW) 55% ± 3.5 (standard deviation). The average LOI value of the concretions was 12% of DW. In the concretions, the concentration of Fe was the highest (average value 4.2 ± 0.8 mol kg-1 DW) and the concentration of Mn was about half of that (2 ± 0.7 mol kg-1 DW). The average concentration of P was 0.7 ± 0.1 mol kg-1 DW, and the concentration of As was 0.0042 ± 0.0007 mol kg-1 DW, respectively. Molar ratios of Fe:Mn in the concretions ranged between 1.1–6.5, whereas those of Fe:P and Fe:As varied less being 5.2–6.7 and 800–1100, respectively. There was a significant positive correlation (Pearson correlation co–efficient) between Fe and P, Fe and As (Fig. 1a, b), and between P and As (r= 0.96, p< 0.0001). The correlation was negative 5 between Mn and Fe (Fig. 1c) as well as between Mn and P (r= –0.9, p< 0.0005) and between Mn and As (r= –0.85, p< 0.005).

3.2 Carbon and ammonium dynamics in incubations The consumption of DOC led to the production of DIC during the incubations (Fig. 2). The added DOC was entirely consumed first at 20 °C and then at 10 °C, whereas at 5 °C about one third of the added DOC was still left after fifteen weeks of incubation. The production of DIC was minor in the incubations without the added C (Fig. 2). Consumption of DOC was low in the autoclaved controls, but yet there was some increase in DIC (data not shown). The increase in DIC was, however, clearly below the values measured from non– sterilized incubations. In general, the consumption of DOC and production of DIC correlated with the + consumption of NH4 in the incubations with C additions (data not shown). The depletion occurred first at 20 °C and then at 10 °C and finally at 5 °C after fifteen weeks of + incubation (Fig. 3a). However, NH4 was consumed below the detection limit only in the + units without the added C (Fig. 3b). Highest concentrations of NH4 were found only from the autoclaved controls with the added C at the end of the incubation (Fig. 3a).

3.3 Effects of labile organic C and temperature on the release of Fe and Mn The released amount of Mn from the concretions to the solution was invariably higher than that of Fe at all incubation temperatures (Fig. 4a), although the concentration of Fe was higher than Mn in the intact concretions (Fig. 1c). The dissolution of Mn was fast at 20 °C and 10 °C in the beginning, but the concentrations decreased at 20 °C at the end of the experiment (Table 2, Fig. 4b). The release rate of Mn was eventually the highest and the concentration reached the highest level at 5 °C at the end of the experiment (Table 2, Fig. 4b). When C was not added, the Mn release rate was 0.009 μmol g-1 DW d-1, and in autoclaved controls with C addition 0.02 μmol g-1 DW d-1. Autoclaving has previously been reported to caus dissolution of organic C, P and Mn from soils and sediment (Tuominen et al. 1994), and it initiated some release of elements and DIC into solution also in this experiment. Fe was released almost steadily into the solution at 5 °C during the experiment (Fig. 4b) whereas at higher temperatures the concentration increased first and decreased later. The proportion of Fe2+ was more than 90% of the total dissolved Fe in all the units. The release of Fe increased after one month of incubation, reaching the highest concentrations later than Mn (Fig. 4b, Table 2). The concentration of Fe dropped at 10 °C and 20 °C (Fig. 4b) simultaneously with the formation of black precipitates (possibly Fe–sulphides) on the bottom of the flasks. A distinct smell of H2S was observed at 10 °C and 20 °C at the end of the experiments. The release rate of Fe remained below 0.002 μmol g-1 DW d-1 in the autoclaved units and the units without C added.

3.4 Effects of labile organic C and temperature on the release of P and As The release of P and As were positively correlated with the release of Fe at 5 °C (r= 0.99, p< 0.001, Fig. 5a, b) but not with Mn (data not shown). Additionally, more P was released (P:Fe 0.19, Fig. 5a) than what was expected on the basis of the P:Fe ratio 0.12 of the solid concretions (Fig. 1a), but at 10 °C and 20 °C the decrease in the concentration of P followed that of Fe. The As:Fe ratio in the solution was close to that measured from the solid concretions (0.7*10-3, Fig. 1b, 5b) at 5 °C. At higher temperatures As:Fe ratio

6 increased (1.2*10-3, Fig. 5b) which was explained mainly by the decreased Fe values at 10 °C and 20 °C. The release of P and As followed the formation of DIC at 5 °C (Fig. 6a, b). In the incubations without the added C, and in the autoclaved controls, the release rate of P stayed below 3*10-3 μmol g-1 DW d-1, and that of As below 0.02*10-3 μmol g-1 DW d-1.

4 Discussion 4.1 Physical structure of concretions Solid metal–rich concretions form a distinct habitat for microbes compared to the surrounding fine–grained sediment. The high water content indicated that the spherical concretions had porous internal structure. The structure is likely to increase specific surface area of the concretion that might be in contact with microbes. Additionally, the high porosity may have allowed water and solutes to move within the concretion and permitted the formation of chemical gradients which may have provided variable environments for microbes as well as precipitation and dissolution of Fe, Mn, P and As within short distances.

4.2 Effects of organic C and ammonium on concretions in anoxic conditions The high organic C content is also characteristic for the spherical concretions. In fact, the measured LOI values were comparable to those found from the surface sediments of the GoF, where LOI values vary from 7–17% up to 23% DW (Lehtoranta 2003; Lehtoranta et al. 2004). Thus, the concretions had captured considerable amounts of organic C during their formation. The organic C is likely co–precipitated and coagulated with or adsorbed on Fe3+–hydroxides (McKnight et al. 1992; Sharma et al. 2010; Eusterhues et al. 2011; Riedel et al. 2013). The release of Fe and Mn and the formation of DIC stayed at low level in the units without C addition suggesting that the bioavailability of the abundant organic C present in the concretions was low for Fe– and Mn–reducing microbes. The organic C was either refractory or it located in the inner parts of the concretions where it may have not been as readily accessible for microbes and their enzymes. Hereby, concretion bottoms may store organic C. Significant amounts of Fe and Mn were released from the concretions to ambient solution only when they were supplied with labile C. The result indicated that bacteria could use the metal oxides present in the concretions as electron acceptors, but needed an external source of C. The dissolution of Fe and Mn at all temperatures indicated that the reduction of Fe and Mn in the concretions was limited by the availability of labile C. In the nutrient–rich GoF the concretion bottoms may receive high quantities of organic C after spring bloom of algae (Heiskanen and Leppänen 1995; Kankaanpää et al. 1997). Sediment bacteria degrade this settling organic matter producing simple C compounds, which may initiate the microbial reduction of metals in the concretions after oxygen has been consumed. Ammonium may adsorb onto sediment particle surfaces (Mackin and Aller 1984). In + the experiment the removal of NH4 was minor in the autoclaved controls, which indicates + that autoclaved concretions did not remove NH4 efficiently. In contrast, there was + decrease in the concentration of NH4 and increase in DIC in the microcosms with the + + added C. The removal of NH4 was likely explained with the assimilation of NH4 into + microbial biomass. However, the added NH4 was completely removed only in the incubations without added organic C. In these units the release of Fe and Mn was minor. + + Anaerobic oxidation of NH4 (i.e. ANAMMOX reaction) which could remove NH4 would 7

- have required nitrite (NO2 , Jetten et al. 1999) but it was not supplied for the microbes (nor - was nitrate, NO3 ). It has been suggested that Mn–oxide reduction could be coupled to + - + + 2+ NH4 oxidation to form NO3 in the following reaction; 4MnO2 + NH4 + 6H –> 4Mn + - NO3 + 5H2O (Hulth et al. 1999), but there is no evidence of the existence of such process in natural environmental conditions (Thamdrup and Dalsgaard 2000; Bartlett et al. 2007). The minor release of dissolved Mn did not support the reaction either.

4.3 Release of Fe and Mn from concretions in anoxic conditions The released amount of Mn exceeded always that of Fe at all incubation temperatures despite the higher Fe content of the concretions. Additionally, the reduction of Mn started earlier than that of Fe which has also been observed in sandy marine sediments (Kristiansen et al. 2002). Several factors may explain the differences in the release of Mn and Fe from the concretions. The spherical concretions have a zoned structure of brown Fe and black Mn oxide layers (Yli–Hemminki et al. 2014), and a Mn coating which may favour the reduction of Mn. The earlier starting point of Mn reduction over Fe in the concretions seems to follow the free energy gain: Mn reduction may have yielded more energy for bacteria. Bacteria may also specialize in Mn reduction (Thamdrup et al. 2000). Furthermore, Mn oxides may participate directly in the oxidation of formed Fe2+ ( Lovley 1991; Postma and Appelo 2000). The reaction would have inhibited the accumulation of Fe into the solution and thus, in part, explain the higher concentrations of Mn in the incubations. Mn oxides can even be reduced by As3+ (Oscarsson et al. 1982; Chen et al. 2006; Zhang et al. 2014). In conclusion, concretion Mn may therefore have a more pronounced role as an electron acceptor, which leads to its accelerated release compared to Fe in the presence of labile C and during occasional reducing conditions in the GoF. By participating in the redox–reactions in this manner in anoxic conditions, as a consequence, Mn may even delay the release of substances sequestered by Fe oxides in the concretions.

4.4 Release of P and As from concretions in anoxic conditions Phosphorus and As were released with Fe from the concretions, and the production of DIC was in line with the accumulation of P and As at 5 °C. The result indicate that P and As were released from the concretions into solution by microbial Fe reduction. The Fe:P ratio 5.2 recorded in the solution was at a lower range limit compared to the ratio in intact concretions. In comparison, the molar ratio Fe:As 1350 in the solution was higher than that in the intact concretions at 5 °C. This indicates favoured release of P from the Fe oxides of the concretions at low temperature.

4.5 Effects of temperature on the release of Fe, Mn, P and As from concretions during anoxia At 5 °C the release of Fe was almost constant and the released Mn remained at its highest level throughout the incubation period. It is evident that at low temperature, Fe and Mn 2- 2- reduction was favoured over SO4 reduction, or that SO4 reduction remained at a level which did not result in noticeable removal of Fe and Mn as sulphides. The elevation of temperature to 10 °C and 20 °C changed the reactions of all the elements studied. The increase of temperature accelerated the release of Fe and Mn in the beginning, but led to the removal of Fe at 10 °C and 20 °C, and of Mn at 20 °C after midpoint of the incubation. It is possible, that the increase of the incubation temperature enhanced microbial Fe and 2- Mn reduction, but then favoured the reduction of SO4 (Lovley and Phillips 1987; Knoblauch and Jørgensen 1999) which was indicated by the noticed smell of H2S at the 8 end of the experiment. The presence of external electron donor, i.e. sulphides, caused Fe and Mn precipitation with sulphides observed as the formation of black precipitates on the bottom of the incubation bottles. In this formation point there was simultaneous removal of Fe and P from the solution. It is known that usually Fe loses its ability to capture P when Fe is precipitated as sulphides (Roden and Edmonds 1997). However, only low concentrations of P were measured from the solution when black precipitates formed at higher temperatures. When Fe concentration is high – as in the concretions – remaining Fe oxides may retain released P, As and Fe2+ in black–coloured suspension (Golterman 1995; Neidhardt et al. 2014). The geochemical environment formed at higher temperatures may have favoured the precipitation of other Fe compounds that are able to capture P. It may be 2+ 3- that the released P was precipitated as Fe –PO4 minerals (i.e. vivianite), as has been suggested by Slomp et al. (2013) to be possible in the Bothian Bay, Baltic Sea. This reaction requires a low concentration of free sulphides and a high Fe concentration (Roden and Edmond, 1997; Slomp et al. 2013). It has also been proposed that P could be associated with Mn in the GoF (Mort et al. 2010; Jilbert and Slomp 2013). Still, experiments have not recorded significant interaction between P and Mn dynamics in the Baltic Sea sediments (Gunnars and Blomqvist 1997). In contrast to P, As tended to stay in the solution with rising temperature, and the Fe:As molar ratio 780 in the solution was lower than in the intact concretions. It is possible that As was bound to Fe complexes which easily released As (Sharma et al. 2010). Dissimilatory As reduction (Ahmann et al. 1997; Akerman et al. 2011) or removal of As from microbial cells (Anderson and Cook 2004) may also contribute to the mobilization of As from concretions. Temperature of the concretion bottoms stays below 4 °C due to density gradient of water and the formation of thermocline and halocline, especially in the western GoF. It has been predicted that climate warming could result in shift towards fresher and warmer surface waters which would lead to weakening of halocline and strengthening of vertical density stratification in thermocline in deeper areas of the GoF (Liblik and Lips 2011). In this scenario, the GoF water column would be mixed to the greater depths during winter. The mixing of water would bring oxygen to the bottom and might protect concretions from metal reduction and decay. On the other hand, warmer surface waters support algal growth 2+ in the nutrient–rich GoF. Microbial processes such as Mn–, Fe– and SO4 – reduction were shown to enhance at the elevated temperatures only in the presence of excess labile C, creating threatening conditions for concretion bottoms.

4.6 Release of P and As from the Fe–Mn concretions vs. sediment effluxes Measurements pointing out the high effluxes of Fe and P from sediments during anoxia have been carried out at low temperatures in the GoF (Lehtoranta and Heiskanen 2003; Pakhomova et al. 2007; Viktorsson et al. 2012). Therefore, the release rates from the concretions measured only at 5 °C were compared to those reported earlier from the sediments of the GoF. The release rates of Fe, Mn and P calculated for the mean amount of concretion deposit per square meter were comparable with those reported earlier from the organic rich accumulation bottoms of the GoF during anoxia (Table 3). The concretions deposits may, thus, form a significant local source of Fe, Mn and P from the bottom to the water besides the intensively studied accumulation bottoms. In the Baltic Sea and the GoF the redox–sensitive Fe–bound P in sediments is considered to control the highly varying concentration of P in bottom water (Lukkari et al. 2009b; Mort et al. 2010; Jilbert et al. 2011; Rydin et al. 2011) and even in the entire water volume. During the 9 anoxia, large amounts of P are released from the sediments counteracting decreases in the external P load to the eutrophic GoF (Pitkänen et al. 2001). However, the Fe:P release ratio 5.2 calculated for the concretions was several times higher than the ratio 0.02–1 calculated for coastal and open sea sediments in The Baltic Sea (Gunnars and Blomqvist 1997), and thus the high Fe:P ratio >2 (Gunnars and Blomqvist 1997; Gunnars et al. 2002) predicts good P–scavenging capacity of the concretion bottoms. Anoxia increases the benthic efflux of As. For example, in the contaminated Balmer Lake in Canada enhanced eutrophication has led to anoxia and subsequent increase in the release of As from sediment (Martin and Pedersen 2002). There are no published studies on the benthic effluxes of As from the GoF, probably due to the relatively low concentrations of As in the sediment (mean 0.19 mmol kg-1, Vallius 2012) which has not encouraged to perform flux measurements. Instead, we compared our experimental release rates of dissolved As (Table 3) to the estimated average accumulation rate of particulate As of 0.6 μmol m-2 d-1 measured from the open GoF from April to October (Leivuori and Vallius 1998). The release rate (range) may be over five times higher than the average sedimentation rate, which suggests that the efflux of As from the concretions may temporarily exceed the flux of particulate As to the GoF. In the GoF, the extent of anoxia may last from days to over a year in the water depths where the concretions were collected (Lehtoranta and Heiskanen 2003; Viktorsson et al. 2012). In the incubations lasting over three months only 0.1, 0.4, 0.1 and 0.009 % of the total Fe, Mn, P and As, respectively, were released from the concretions to the solution at 5 °C. In comparison, over one third of the total Fe and Mn pool has been reported to be released in sandy coastal marine sediment surface within a month (Kristiansen et al. 2002). Additionally, a significant pool of Mn has been mobilized within days reaching the maximum release rate already in ten days in hypoxic marine Baltimor harbour–sediment surface (Riedel et al. 1999). In the coastal sediments the major release of P bound to Fe has occurred within few months in autumn (Jensen et al. 1995). In the concretions the released proportion of As (0.009%) formed almost an insignificant fraction of the total As, whereas even 28% of the total As in the anoxic sediment has been mobilized in Aberjona watershed in the U.S.A. (Ahmann et al. 1997). Based on our experiment, it is unlikely that the bottom water anoxia lasting for months would result in major decay of the pools of Fe and Mn oxides, and concomitant release of P and As from concretions similarly as the bottom surface sediments of the GoF. The resistance of concretions towards anoxia may be better than that of sediments due to the physical form and geochemical properties of the concretions. Additionally, the availability of labile C may be lower for the Fe and Mn reducing microbes in the concretions than for those living in the organic rich accumulation bottoms of the GoF.

4.7 Geochemical characteristics of concretions in the Baltic Sea The geochemical characteristics reported in the literature on the concretions from different environments were compared to each other to find out their similarities and differences, and to reveal their possible sensitivity towards anoxia. The formation rate of the concretions in the Baltic Sea is estimated to be 14 μm a-1 (Grigoriev et al. 2013), which is orders of magnitude higher than the value of 0.01 μm a-1 for the oceanic concretions (Glasby et al. 1982). Already the short geological history of the Baltic Sea reveals that concretions are rather young (max. 3000 years, Ingri 1985) and their mean age can be between 670– 850 years (Grigoriev et al. 2013). The concretions in the Baltic Sea may, thus, capture Fe, Mn, P, and As, and even organic C much faster than their Oceanic 10 counterparts when the redox–conditions allow their growth. The alternating layers of Fe and Mn observed from the GoF have also been found from the soil concretions in Greece (Gasparatos et al. 2005), and in Fe–Mn crusts in the south–west Baltic Sea (Marcus et al. 2004). In concretions the molar ratio between Fe and Mn varies depending on the environment they originate from (Fig. 7a, Table A1, Electronic Supplementary Material). In general, the concretions of the sub–basins of the Baltic Sea accumulate more Fe than those found from the Oceans and more Mn than soil concretions. Hence, the concretions from the sub–basins of the Baltic Sea fall into an intermediate group between fresh–water and deep–sea concretions (see also Ingri 1985). The correlation between Fe and Mn in the concretions is negative in the Baltic Sea sub basins (r= -0.9, p< 0.0005), and weakly (i.e. not significantly) negative in the Oceans (r= -0.2, p< 0.4) and in lakes (r= -0.3, p< 0.5), but in soils the correlation is weakly positive (soil r= 0.35, p< 0.07, Fig. 7a). The contrasting correlations indicate that the concretions of the brackish marine and true marine systems seem to have higher tendency to retain Mn over Fe than the lake and soil systems. This might be related to the availability of Fe and Mn, but also to different microbial redox– reactions which concretions encounter in different geochemical environments. Despite the evident differences, the comparisons suggest that the concretions of the Baltic Sea, oceans, lakes, and soils have high concentrations of redox–sensitive Fe and Mn, and thus, share same basic geochemical characteristics which may favor microbial reduction processes affecting the cycling of elements in anoxia. Concretions have a great regional variation in their Fe:P ratio (Fig. 8a). The Baltic Sea concretions have lower Fe:P ratio compared to the oceanic ones, probably due to the higher abundance of P in the sediments of the Baltic Sea and slightly higher concentration of Fe in concretions (Fig. 8a). On the basis of the correlations from the sub–basins of the Baltic Sea one mole of Fe is able to bind around 0.1 moles of P in the concretions (Fig. 8a). The much higher concentrations of Fe, Mn, P, and As in the concretions than in sediment or in settling matter indicate that concretions sequester significant quantities of these elements from their surroundings (Figures 7a, b, 8a, b). Following generalizations are made based on the values calculated from the accumulation of particulate matter from April to October (Leivuori and Vallius 1998). For deeper analyses one has to take into consideration temporal and spatial variation of sedimentation (Blomqvist and Larron 1994). The correlations between Fe and Mn in the sediments is positive (r= 0.85, p< 0.0005) and weakly positive (i.e. not significant) in the settling particulate matter (r= 0.3, p< 0.4) of the GoF, in contrast to the Baltic concretions (Fig. 7b). Additionally, Fe:Mn ratios in the concretions show that they accumulate especially Mn in comparison to the sediments. Only the settling particulate matter collected from the deep waters seem to have similar tendency to accumulate Mn as the concretions (Fig. 7b). In the GoF the sediments as well as settling matter contains much less P compared to the concretions (Fig. 8b). The weak (i.e. not significant) correlation found between sediment Fe and P (r= 0.16, p< 0.7, Fig. 8b) may be explained by the high concentration of Fe–sulphides in the sediments of the deep accumulation bottoms having poor capacity to bind P (Gunnars and Blomqvist 1997). The concentration of As vary only a little in the settling matter (Fig. 9), and in the sediments (Vallius 2012). In the concretions As levels are commonly over ten times higher than in the surface sediments of the GoF (Fig. 9, also Zhamoida et al. 2007). The high industrial discharges are likely to explain the high As levels in the Gulf of Bothnia (Ingri 1985; Leivuori and Niemistö 1995 and references therein). 11

5 Conclusions This study suggests that in anoxic conditions with the presence of labile organic C the spherical concretions rich in Fe and Mn dissolve through microbial reactions resulting in the release of Mn as well as simultaneous release of Fe, P and As. Additionally, the + spherical concretions may participate in the cycling of N by removing NH4 . The experimental release rates of P calculated for the concretion deposits indicate that concretions may form a temporal source of P in anoxia, which may have ecological significance in the GoF. It is, however, evident that the concretions maintain their binding capacity for P and As longer than the organic rich and fine–grained surface sediments in anoxic conditions. Our comparisons of the concretions from the different environments demonstrated that through the high concentrations of redox–sensitive metals, they may become a target for microbial processes in fluctuating oxygen conditions. Thus, the concretions from various regions may respond differently to anoxia by releasing metals and other elements bound to concretions, and hence affect the current geochemical conditions in their surroundings.

Acknowledgements This work was financially supported by Maj and Tor Nessling Foundation, K.H. Renlund Foundation and Maa– ja vesitekniikan tuki ry. The authors wish to thank Suvi Tyynelä for her contribution to the laboratory work, and R/V Muikku and Anna Downey for the sampling of concretions. Conflict of interest: The authors declare that they have no conflict of interest.

References Ahmann D, Krumholz LR, Hemond HF, Lovley DR, Morel FMM (1997) Microbial mobilization of arsenic from sediments of the Aberjona Watershed. Environ Sci Technol 31:2923–2930 Akerman NH, Price RE, Pichler T, Amend JP (2011) Energy sources for chemolithotrophs in an arsenic– and iron–rich shallow–sea hydrothermal system. Geobiology 9:436– 445 Anderson CF, Cook GM (2004) Isolation and characterization of arsenate–reducing bacteria from arsenic–contaminated sites in New Zealand. Curr Microbiol 48:341– 347 Arif M, Blinov LN (2004) Investigation of concretions from the Baltic Sea. Glass Phys Chem 30:359–361 Arshad MA, Arnaud RJ (1980) Occurrence and characteristics of ferromanganiferous concretions in some Saskatchewan soils. Can J Soil Sci 60:685–695 Bartlett R, Mortimer RJG, Morris KM (2007) The biogeochemistry of a manganese–rich Scottish sea loch: Implications for the study of anoxic nitrification. Cont Shelf Res 27:1501–1509 Baturin GN (2009) Geochemistry of ferromanganese nodules in the Gulf of Finland, Baltic Sea. Lithol Miner Resour 44:411–426 Baturin GN, Dubinchuk VT (2009) Composition of ferromanganese nodules from Riga Bay (Baltic Sea). Oceanology 49:111–120 Bischoff JL, Piper DZ, Leong K (1981) The aluminosilicate fraction of North Pacific manganese nodules. Geochim Cosmochim Ac 45:2047–2063

Blomqvist S, Larsson U (1994) Detrital bedrock elements as tracers of settling resuspended particulate matter in a coastal area of the Baltic Sea. Limnol Oceanogr 39:880–896 Chen Z, Kim K–W, Zhu Y–G, McLaren R, Liu F, He J–Z (2006) Adsorption (AsIII,V) and oxidation (AsIII) of arsenic by pedogenic Fe–Mn nodules. Geoderma 136:566–572 Ehrlich HL (1968) Bacteriology of manganese nodules II. Manganese oxidation by cell– free extract from a manganese nodule bacterium. Appl Microbiol 16:197–202 Eusterhues K, Rennert T, Knicker H, Kogel–Knabner I, Totsche KU, Schwertmann U (2011) Fractionation of organic matter due to reaction with ferrihydrite: coprecipitation versus adsorption. Environ Sci Technol 45:527–533 Farquhar ML, Charnock JM, Livens FR, Vaughan DJ (2002) Mechanisms of arsenic uptake from aqueous solution by interaction with goethite, lepidocrocite, mackinawite,and pyrite: An X–ray absorption spectroscopy study. Environ Sci Technol 36:1757–1762 Gasparatos D, Tarenidis D, Haidouti C, Oikonomou G (2005) Microscopic structure of soil Fe–Mn nodules: environmental implication. Environ Chem Lett 2:175–178 Glasby GP (1973) Mechanisms of enrichment of rarer elements in marine manganese nodules. Mar Chem 1:105–125 Glasby GP, Emelyanov EM, Zhamoida VA, Baturin GN, Leipe T, Bahlo R, Bonacker P (1997) Environments of formation of ferromanganese concretions in the Baltic sea: a critical review. In: Nicholson K, Hein JR, Bühn B, Dasgupta S (eds) Manganese mineralization: Geochemistry and mineralogy of terrestrial and marine deposits. Geological society Special publication 119:213–237 Glasby GP, Stoffers P, Sioulas A, Thijssen T, Friedrich G (1982) Manganese nodule formation in the Pacific Ocean: a general theory. Geo–Mar Lett 2:47–53 Ghiorse WC, Hirsch P (1982) Isolation and properties of ferromanganese–depositing buddimg bacteria from Baltic sea ferromanganese concretions. Appl Environ Microb 43:1464–1472 Grigoriev AG, Zhamoida VA, Gruzdov KA, Krymsky Grigoriev RSh (2013) Age and growth rates of ferromanganese concretions from the Gulf of Finland derived from 210Pb measurements. Oceanology 53:345–351 Golterman HL (1995) The role of the iron hydroxide–phosphate–sulphide system in the phosphate exchange between sediments and overlying water. Hydrobiologia 297:43– 54 Gunnars A, Blomqvist S (1997) Phosphate exchange across the sediment–water interface when shifting from anoxic to oxic conditions – an experimental comparison of freshwater and brackish–marine systems. Biogeochemistry 37:203–226 Gunnars A, Blomqvist S, Johansson P, Andersson C (2002) Formation of Fe(III) oxyhydroxide colloids in freshwater and brackish seawater, with incorporation of phosphate and calcium. Geochim Cosmochim Ac 66:745–758 Haahti H, Kangas P (2006) State of the Gulf of Finland in 2004. MERI– Report series of the Finnish Institute of Marine Research. No 55. Helsinki, Finland Hansson M, Andersson L, Axe P (2011) Areal extent and volume of anoxia and hypoxia in the Baltic Sea, 1960–2011. SMHI Swedish Meteorological and Hydrological Institute, Report Oceanography No. 42, Norrköping, Sweden, 76 pp He J, Zhang L, Jin S, Zhu Y (2008) Bacterial communities inside and surrounding soil iron–manganese nodules. Geomicrobiol J 25:14–24

Heiskanen AS, Leppänen JM (1995) Estimation of export production in the coastal Baltic Sea: Effect of resuspension and microbial decomposition on sedimentation measurements. Hydrobiologia 316:211–224 Hohmann C, Winkler E, Morin G, Kappler A (2010) Anaerobic Fe(II)–oxidizing bacteria show As resistance and immobilize As during Fe(III) mineral precipitation. Environ Sci Technol 44:94–101 Hulth S, RC Aller, F Gilbert (1999) Coupled anoxic nitrification/manganese reduction in marine sediments. Geochim Cosmochim Ac 63:49–66 Ingri J (1985) Geochemistry of ferromanganese concretions and associated sediments in the Gulf of Finland. Dissertation. Luleå, Sweden: University of Technology Jensen HS, Mortensen PB, Andersen FØ, Rasmussen E, Jensen A (1995) Phosphorus cycling in a coastal marine sediment, Aarhus Bay, Denmark. Limnol Oceanogr 40:908–917 Jetten MSM, Strous M, van de Pas–Schoonen KT, Schalk J, van Dongen UGJM, van de Graaf AA, Logemann S, Muyzer G, van Loosdrecht MCM, Kuenen JG (1999) The anaerobic oxidation of ammonium. FEMS Microbiol Rev 22:421–437 Jilbert T, Slomp CP (2013) Iron and manganese shuttles control the formation of authigenic phosphorus minerals in the euxinic basins of the Baltic Sea. Geochim Cosmochim Ac 107:155–169 Jilbert T, Slomp CP, Gustafsson BG, Boer W (2011) Beyond the Fe–P–redox connection: preferential regeneration of phosphorus from organic matter as a key control on Baltic Sea nutrient cycles. Biogeosciences 8:1699–1720 Kankaanpää H, Korhonen M, Heiskanen A–S, Suortti A–M (1997) Seasonal sedimentation of organic matter and contaminants in the Gulf of Finland. Boreal Environ Res 2:257–274 Knoblauch C, Jørgensen BB (1999) Effect of temperature on sulphate reduction, growth rate and growth yield in five psychrophilic sulphate–reducing bacteria from Arctic sediments. Environ Microbiol 1:457–467 Kristiansen KD, Kristensen E, Jensen MH (2002) The influence of water column hypoxia on the behaviour of manganese and iron in sandy coastal marine sediment. Estuar Coast Shelf S 55:645–654 Lehtoranta J (2003) Dynamics of sediment phosphorus in the brackish Gulf of Finland. Monographs of the Boreal Environ Res 24. ISBN 952–11–1400–2 (PDF) Lehtoranta J, Heiskanen A–S (2003) Dissolved iron:phosphate ratio as an indicator of phosphate release to oxic water of the inner and outer coastal Baltic Sea. Hydrobiologia 492:69–84 Lehtoranta J, Heiskanen A–S, Pitkänen H (2004) Particulate N and P characterizing the fate of nutrients along the estuarine gradient of the River Neva (Baltic Sea). Estuar Coast Shelf S 61:275–287 Lehtoranta J, Pitkänen H (2003) Binding of phosphate in sediment accumulation areas of the eastern Gulf of Finland, Baltic Sea. Hydrobiologia 492:55–67 Leivuori M (1998) Heavy metal contamination in surface sediments in the Gulf of Finland and comparison with the Gulf of Bothnia. Chemosphere 36:43–59 Leivuori M, Jokšas K, Seisuma Z, Petersell V, Larsen B, Pedersen B, Floderus S (2000) Distribution of heavy metals in sediments of the Gulf of Riga, the Baltic Sea. Boreal Environ Res 5:165–185 Leivuori M, Niemistö L (1993) Trace metals in the sediments of the Gulf of Bothnia. Aqua Fennica 23:89–100 14

Leivuori M, Niemistö L (1995) Sedimentation of trace metals in the Gulf of Bothnia. Chemosphere 31:3839–3856 Leivuori M, Vallius H (1998) A case study of seasonal variation in the chemical compositions of accumulating suspended sediments in the central Gulf of Finland. Chemosphere 36:2417–2435 Leppäranta M, Myrberg K (2009) Physical oceanography of the Baltic Sea. Springer– Praxis books in geophysical sciences. ISBN 978–3–540–79702–9 Liblik T and Lips U (2011) Characteristics and variability of the vertical thermohaline structure in the Gulf of Finlanf in summer. Boreal Environ Res 16 suppl. A:73–83 Lovley DR (1991) Dissimilatory Fe(III) and Mn(IV) reduction. Microbiol Rev 55:259–287 Lovley DR, Phillips EJP (1987) Competitive mechanisms for inhibition of sulfate reduction and methane production in the zone of ferric iron reduction in sediments. Appl Environ Microb 53:2636–2641 Lukkari K, Leivuori M, Kotilainen A (2009a) The chemical character and behaviour of phosphorus in poorly oxygenated sediments from open sea to organic–rich inner bay in the Baltic Sea. Biogeochemistry 96:25–48 Lukkari K, Leivuori M, Vallius H, Kotilainen A (2009b) The chemical character and burial of phosphorus in shallow coastal sediments in the northeastern Baltic Sea. Biogeochemistry 94:141–162 Mackin JE and Aller RC. 1984. Ammonium adsorption in marine sediments. Limnol Oceanogr 29:250–257 Marcus MA, Manceau A, Kersten M (2004) Mn, Fe, Zn and As speciation in a fast– growing ferromanganese marine nodule. Geochim Cosmochim Ac 68:3125–3136 Martin AJ, Pedersen TF (2002) Seasonal and interannual mobility of arsenic in a lake impacted by metal mining. Environ Sci Technol 36:1516–1523 McKnight DM, Bencala KE, Zellweger GW, Aiken GR, Feder GL, Thorn KA (1992) Sorption of dissolved organic carbon by hydrous aluminum and iron oxides occurring at the confluence of Deer Creek with the Snake River, Summit County, Colorado. Environ Sci Technol 26:1388–1396 Mort HP, Slomp CP, Gustafsson BG, Andersen TJ (2010) Phosphorus recycling and burial in Baltic Sea sediments with contrasting redox conditions. Geochim Cosmochim Ac 74:1350–1362 Müller B, Granina L, Schaller T, Ulrich A, Wehrli B (2002) P, As, Sb, Mo, and other elements in sedimentary Fe/Mn layers of Lake Baikal. Environ Sci Technol 36:411– 420 Neidhardt H, Berner ZA, Freikowski D, Biswas A, Majumder S, Winter J, Gallert C, Chatterjee D, Norra S (2014) Organic carbon induced mobilization of iron and manganese in a West Bengal aquifer and muted response of groundwater arsenic concentrations. Chem Geol 367:51–62 Oscarsson DW, Huang PM, Hammer VT (1983) Oxidation and sorption of arsenite by manganese dioxide as influenced by surface coatings of iron and aluminium oxides and calcium carbonate. Water Air Soil Poll 20:233–244 Pakhomova SV, Hall POJ, Kononets MY, Rozanov AG, Tengberg A, Vershinin AV (2007) Fluxes of iron and manganese across the sediment–water interface under various redox conditions. Mar Chem 197:319–331 Pitkänen H, Lehtoranta J, Peltonen H, Laine A, Kotta J, Kotta I, Moskalento P, Mäkinen A et al. (2003) Benthic release of phosphorus and its relation to environmental

conditions in the estuarial Gulf of Finland, Baltic Sea in the early 2000s. Proceedings of the Estonian Academy of Sciences, Biology and Ecology 52:173–193 Pitkänen H, Lehtoranta J, Räike A (2001) Internal nutrient fluxes counteract decreases in external load: the case of the estuarial eastern Gulf of Finland, Baltic Sea. Ambio 30:195–201 Plugge CM (2005) Anoxic media design, preparation and considerations. Method Enzymol 397:3–16 Postma D, Appelo CAJ (2000) Reduction of Mn–oxides by ferrous iron in a flow system: Column experiment and reactive transport modelling. Geochim Cosmochim Ac 64:1237–1247 Riedel GF, Sanders JG, Osman RW (1999) Biogeochemical control on the flux of trace elements from estuarine sediments: effects of seasonal and short–term hypoxia. Mar Environ Res 47:349–372. Riedel T, Zak D, Biester H, Dittmar T (2013) Iron traps terrestrially derived dissolved organic matter at redox interfaces. PNAS 110:10101–10105 Roden EE, Edmonds JW (1997) Phosphate mobilization in iron–rich anaerobic sediments: Microbial Fe(III) oxide reduction versus iron–sulfide formation. Arch Hydrobiol 39:347–378 Rona P, Lenoble J (2004) Marine mineral resources: scientific advances and economic perspectives. United Nations Division for Ocean Affairs and the Law of the Sea, Office of Legal Affairs, and the International Seabed Authority 62 Rydin E, Malmaeus JM, Karlsson OM, Jonsson P (2011) Phosphorus release from coastal Baltic Sea sediments as estimated from sediment profiles. Estuar Coast Shelf S 92:111–117 Savchuk OP (2000) Studies of the assimilation capacity and effects of nutrient load reductions in the eastern Gulf of Finland with geochemical model. Boreal Environ Res 5:147–163. Sharma P, Ofner J, Kappler A (2010) Formation of binary and ternary colloids and dissolved complexes of organic matter, Fe and As. Environ Sci Technol 44:4479– 4485 Slomp CP, Mort HP, Jilbert T, Reed DC, Gustafsson BG, Wolthers M (2013) Coupled dynamics of iron and phosphorus in sediments of an oligotrophic coastal basin and the impact of anaerobic oxidation of methane. PLOS One 8:e62386. DOI: 10.1371/journal.pone.0062386 Suess E, Djafari D (1977) Trace metal distribution in Baltic Sea ferromanganese concretions: inferences on accretion rates. Earth Planet Sc Lett 35:49–54 Stookey LL (1970) Ferrozine – A new spectrophotometric reagent for Iron. Anal Chem 42:779–781 Szymanski W, Skiba M (2013) Distribution, morphology, and chemical composition of Fe–Mn nodules in albeluvisols of the Carpathian Foothills, Poland. Pedosphere 23:445–454 Tan W–F, Liu F, Li Y–H, Hu H–Q, Huang Q–Y (2006) Elemental composition and geochemical characteristics of iron–manganese nodules in main soils of China. Pedosphere 16:72–81 Tebo BM, Emerson S (1986) Microbial manganese(II) oxidation in the marine environment: A quantitative study. Biogeochemistry 2:149–161 Thamdrup B, Dalsgaard T (2000) The fate of ammonium in anoxic manganese oxide–rich marine sediment. Geochim Cosmochim Ac 64:4157–4164 16

Thamdrup B, Rosselló–Mora R, Amann R (2000) Microbial manganese and sulfate reduction in Black Sea shelf sediments. Appl Environ Microb 66:2888–2897 Tuominen L, Kairesalo T, Hartikainen H (1994) Comparison of methods for inhibiting bacterial activity in sediment. Appl Environ Microb 60:3454–3457 Vallius, H (2012) Arsenic and heavy metal distribution in the bottom sediments of the Gulf of Finland through the last decades. Baltica 25:23–32 Vallius H, Leivuori M (1999) The distribution of heavy metals and arsenic in recent sediments in the Gulf of Finland. Boreal Environ Res 4:19–29 Vallius H, Ryabchuk D, Kotilainen A (2007) Distribution of heavy metals and arsenic in soft surface sediments of the coastal area off Kotka, Northeastern Gulf of Finland, Baltic Sea. In: Vallius H (ed.) Holocene sedimentary environment and sediment geochemistry of the Eastern Gulf of Finland, Baltic Sea. Geological Survey of Finland, Special Paper 45:33–48 Vallius H, Zhamoida V, Kotilainen A, Ryabchuk D (2011) Seafloor desertification – A future scenario for the Gulf of Finland? Central and Eastern European Development Studies 365–372. DOI: 10.1007/978–3–642–17220–5_17 Viktorsson L, Almroth–Rosell E, Tengberg A, Vankevich R, Neelov I, Isaev A, Kravtsov V, Hall POJ (2012) Benthic phosphorus dynamics in the Gulf of Finland, Baltic Sea. Aquat Geochem 18:543–564 Viktorsson L, Ekeroth N, Nilsson M, Kononets M, Hall POJ (2013) Phosphorus recycling in sediments of the central Baltic Sea. Biogeosciences 10:3901–3916 Winterhalter B (1966) Pohjanlahden ja Suomenlahden rauta–mangaani–saostumista (Ferromanganese concretions of the Gulf of Bothnia and the Gulf of Finland). Geoteknillisiä julkaisuja N:o 69 (in Finnish) Wu Y–H, Liao L, Wang C–S, Ma W–L, Meng F–X, Wu M, Xu X–W (2013) A comparison of microbial communities in deep–sea polymetallic nodules and the surrounding sediments in the Pacific Ocean. Deep–Sea Res I 79:40–49 Yli–Hemminki P, Jørgensen KS, Lehtoranta J (2014) Iron–manganese concretions sustaining microbial life in the Baltic Sea: The structure of the bacterial community and enrichments in metal–oxidizing conditions. Geomicrobiol J 31:263–275 Zhamoida V, Grigoriev A, Gruzdov K, Ryabchuk D (2007) The influence of ferromanganese concretions–forming processes in the eastern Gulf of Finland on the marine environment. In: Vallius H (ed.) Holocene sedimentary environment and sediment geochemistry of the Eastern Gulf of Finland, Baltic Sea. Geological Survey of Finland, Special Paper 45:21–31 Zhang G, Liu F, Liu H, Qu J, Liu R (2014) Respective role of Fe and Mn oxide contents for arsenic sorption in iron and manganese binary oxide: An X–ray absorption spectroscopy investigation. Environ Sci Technol 48:10316–10322

Table 1 The sampling sites of the Fe–Mn concretions used in this experiment

The Gulf of Location Water Shape and size of Finland, depth concretions Finnish–Estonian (m) EEZ area, July 2010 station (st.) 1 Lat 59°57´43 N, 62 Spherical, 0.5–2 cm in st. 12 Long 25°12´76 E diameter st. 17 Lat 59°57´24 N, 54 Long 24°59´58 E Lat 59°57´59 N, 51 Long 24°51´05 E

Table 2 Release rates of Fe, Mn, P and As from the Gulf of Finland concretions during anoxic incubations at different temperatures

Temperature 5 °C 10 °C 20 °C Element μmol g-1 DW concretion d-1 Fe 0.01–0.12 0.02–0.61 0.02–0.22 Mn 0.002–0.42 0.003–0.24 0.14–0.31 P 0.01–0.02 0.003–0.05 0.003–0.05 As 0.02*10-3 –0.2*10-3 0.03*10-3 – 0.01*10-3 – 0.2*10-3 0.3*10-3

Table 3 Estimated release rates of Fe, Mn, P and As during anoxia from the concretion bottoms of the GoF (values derived from the experiment at 5°C) compared to previously published fluxes of Fe, Mn and P out of the sediments of the GoF and other marine locations (Jensen et al. 1995; Riedel et al. 1999; Lehtoranta 2003; Lehtoranta and Heiskanen 2003; Pakhomova et al. 2007; Lukkari et al. 2009a,b; Viktorsson et al. 2012; Viktorsson et al. 2013)

This study Earlier published studies Element Dissolution rate Estimated flux out Reference Location of the range of the sediment reference μmol m-2 d-1 μmol m-2 d-1 sediment (15 kg concretions m-2 in the GoF on average) Fe 84–1500 100–1000 Pakhomova et GoF al. 2007 Fe 84–1500 90–107 Lehtoranta & GoF Heiskanen 2003 18

Mn 30–6300 500–4450 Pakhomova et GoF al. 2007 P 190–550 540 (max) Jensen et al. Aarhus Bay, 1995 Denmark P 190–550 161 (min), 1033 Lehtoranta 2003 GoF (max), 420 (average) P 190–550 161–1065 Lehtoranta & GoF Heiskanen 2003 P 190–550 198–1774, Lukkari et al. GoF 46–1065 2009a, b P 190–550 1250 ± 560 Viktorsson et al. GoF 2012 P 190–550 376 ± 214 Viktorsson et al. Baltic proper 2013 As 0.2–3.3 1.6 (max) Riedel et al. Baltimor 1999 Harbour sediment U.S.A.

Table A1. Locations and references to published studies on Fe–Mn concretions, sediments and settling particulate matter used in the comparisons to the concretions from this study in Figs. 7-9

Reference Location Depth (m) n of Characteristics values Concretions rev. Ingri 1985 Gulf of Finland 1 Zhamoida et al. GoF 50 7 2007 Ingri 1985 Gulf of Bothnia 48 35 spheroidal, mean discoidal, values crust Baturin and Gulf of Riga 45–55 4 Dubinchuk 2009 rev. Glasby et al. GoR 1 1997 rev. Ingri 1985 GoR 1 rev. Ingri 1985 Western Baltic 2 Sea Suess and Djafari WBS 25 1 1977 rev. Arif and Blinov Pacific, Indian, 3 2004 Atlantic oceans rev. Baturin and Oceans 1 Dubinchuk 2009 Bischoff et al. 1981 North Pacific 5000 9 ocean Glasby 1973 Indian ocean, 4042, 1280, 5 Campbell 806, 2405, mean Plateau, Blake 190 values Plateau, Gulf of Aden, Loch Fyne rev. Ingri 1985 Pacific, Atlantic 3 and Indian mean oceans values Rona and Lenoble Oceans 1 2004 Wu et al. 2012 Pacific ocean 5178–5513 3 rev. by Ingri 1985 Fresh water: 7 Sweden, mean Finland, U.S.A. values Arshad and Arnaud Soil in 0.8 19 podzol, 1980 Saskatchewan glaysol, in Canada luvisol He et al. 2008 Soil in Wuhan 0.2–0.4 1 subacid orthic 20

in Central agrudalf China Szymanski and Soil in 0.2–1.8 8 albeluvisols Skiba 2013 Carpathian Foothills in Poland Tan et al. 2006 Soil in Guangxi, 0.25–1 1 Hunan, Hubei and Shandong in China Sediment (0-1 cm) Leivuori 1998 Gulf of Finland < 60 < 5 Lukkari et al. 2009a, GoF 11 b Pakhomova et al. GoF 43–86 9 aleuritic– 2007 pelitic mud with Fe–Mn concretions Vallius 2012 GoF 11 Vallius and Leivuori GoF 1 muddy silt and 1999 clay Vallius et al. 2007 GoF 14 Zhamoida et al. GoF 1 2007 Ingri 1985 Gulf of Bothnia 4 oxidized surface sediment, weak halocline Leivuori 1998 GoB 1 Leivuori and GoB 43–68 4 Salinity 0–7 Niemistö 1993, ‰, well 1995 oxygenated bottom Ingri 1985 Bothnian Sea 1 Leivuori and BS 4 Niemistö 1993, 1995 Leivuori 1998 BS 1 Leivuori et al. 2000 Gulf of Riga 1–50 1 Muller et al. 2002 Lake Baikal 3 Settling particulate matter Leivuori and Vallius Gulf of Finland 73 12 1998 21

Figures with captions

Fig. 2 Trends in concentrations of dissolved organic C (DOC, solid line) and dissolved inorganic C (DIC, dashed line) with C and without C (no line) in anoxic incubations of the Gulf of Finland concretions at 5 °C, 10 °C and 20 °C

Fig. 1 Concentrations of elements (Fe, Mn, P, As) and correlations between a. Fe and P, b. Fe and As, c. Fe and Mn in intact concretions from the Gulf of Finland, Baltic Sea

+ Fig. 3 Concentrations of ammonium (NH4 ) in the overlying water a. with C, b. without C in anoxic incubations of the Gulf of Finland concretions at 5 °C, 10 °C and 20 °C. AC represents autoclaved control (no line)

Fig. 4 Released concentrations of Fe and Mn to the overlying water from the anoxic Fig. 5 Correlations between the released incubations of the Gulf of Finland amounts of a. Fe and P, b. Fe and As at 5 concretions with added C at 5 °C, 10 °C and °C, 10 °C, and 20 °C to the overlying water 20 °C. a. Molar ratios of released Fe and Mn from the anoxic incubations of the Gulf of b. Release of dissolved Mn and Fe2+ during Finland concretions with added C fifteen weeks. The solid line denotes subsampling from the same unit

Fig. 6 Relationship between formed DIC and released a. P, b. As in 5 °C, 10 °C and 20 °C. Arrows denote the depletion points for dissolved organic carbon and formation of black precipitates in the test units

Fig.7 Correlations between Fe and Mn a. in Fig. 8 Correlations between Fe and P a. in concretions from this study compared to the concretions from this study compared to the ones originating from The Gulf of Finland ones found from The Gulf of Finland (GoF), Gulf of Bothnia (GoB), Gulf of Riga (GoF), Gulf of Bothnia (GoB), Gulf of Riga (GoR), western Baltic Sea (WBS), Oceans, (GoR), western Baltic Sea (WBS), Oceans, lakes and soils, b. from surrounding lakes and soils, b. from sediments and sediments and settling particulate matter in settling particulate matter in the GoF and the different sub basins of the Baltic Sea GoB (Arshad and Arnaud 1980; Bischoff et (Glasby 1973; Suess and Djafari, 1977; al. 1981; Ingri 1985; Leivuori and Vallius Arshad and Arnaud 1980; Bischoff et al. 1998; Muller et al. 2002; Tan et al. 2006; 1981; Ingri 1985; Leivuori and Niemistö Zhamoida et al. 2007; Baturin and 1993, 1995, Glasby et al. 1997; Leivuori Dubinchuk 2009; Lukkari et al. 2009a, b; 1998; Leivuori and Vallius 1998; Leivuori Szymanski and Skiba 2013; Wu et al. 2013. et al. 2000; Muller et al. 2002; Arif and See Table A1, Electronic Supplementary Blinov 2004; Rona and Lenoble 2004; Tan Material) et al. 2006; Pakhomova et al. 2007; Zhamoida et al. 2007; He et al. 2008; Baturin and Dubinchuk 2009; Lukkari et al. 2009a, b; Wu et al. 2013; Szymanski and Skiba 2013. See Table A1, Electronic Supplementary Material)

Fig. 9 Fe and As concentrations in intact concretions from this study, the ones originating from The Gulf of Finland (GoF), Gulf of Bothnia (GoB), Gulf of Riga (GoR), Oceans and lakes, and surrounding sediments and settling particulate matter in the GoF and GoB (Glasby 1973; Ingri 1985; Leivuori and Niemistö 1993, 1995; Leivuori 1998; Leivuori and Vallius 1998; Vallius and Leivuori 1999; Vallius et al. 2007; Zhamoida et al. 2007; Baturin and Dubinchuk 2009; Vallius 2012. See Table A1, Electronic Supplementary Material)

III

Degradation of crude oil and PAHs in the iron–manganese concretion and sediment bottoms from the northern Baltic Sea

Anna Reunamo*, Pirjo Yli–Hemminki, Jari Nuutinen, Jouni Lehtoranta, Kirsten S. Jørgensen

Finnish Environment Institute, P.O. Box 140, FI–00251, Helsinki, Finland

*Corresponding Author: Anna Reunamo Finnish Environment Institute P.O. Box 140 FIN–00251 Helsinki, Finland Tel. +358 295 251 557 E–mail address: [email protected]

Running title: Crude oil degradation in Fe–Mn concretions

Abstract The ability of different sea bottom types to recover from oil contamination under oxic and anoxic conditions is poorly known in the brackish Baltic Sea. We carried out microcosm experiments to examine the capacity of iron–manganese concretions and sediment to remove petroleum compounds at 10 °C. The biological degradation potential of PAHs was indicated by the increase in the copy numbers of PAH–degradation genes up to 2.2 × 106 gene copies g-1 DW. The disappearance of crude oil was 35−81% and that of PAHs around 96% with little difference in the removal between the concretions and sediment. Bacterial community analysis revealed that the bacterial sequences obtained from concretions were affiliated to groups typical for concretions and metal rich environments even after the oil exposure, whereas sediment sequences were similar to those originating from sediments and oil–contaminated sites. Our study suggests that concretions and bottom sediment maintain rich, distinct bacterial communities under oil exposure having potential to remove petroleum compounds in oxic and anoxic conditions.

Keywords: bacterial communities, biodegradation, Fe–Mn concretions, naphthalene mineralization, PAH–RHDα–gene, quantitative PCR, sediment

1. Introduction The Baltic Sea is under a pressure of dense traffic and approximately 160 million tons of oil is transported via the Gulf of Finland annually (Helcom 2012). Constant exposure to small oil spills from ships may enrich petroleum hydrocarbon–degrading groups within the bacterial communities. Microbial diversity and the potential of microbes to degrade oil compounds is widely studied in soils and in the oceans (reviewed by Leahy and Colwell 1990). However, the ability of the microbial community to degrade petroleum hydrocarbons is poorly studied in cold and brackish systems such as the Baltic Sea, where the risk of a major oil accident exists due to increasing volume of oil transport (Helcom 2012). In the oceans Gammaproteobacteria are most often responsible for the degradation process for example in the Gulf of Mexico (Hazen et al. 2010, Gutierrez et al. 2013), The North Sea (Brakstad and Lodeng 2005) and the Mediterranean (Cappello et al. 2007). In the brackish Baltic Sea, on the other hand, the dominant bacterial phyla responsible for the degradation of oil seem to differ along the salinity gradient: in lower salinity Betaproteobacteria and Actinobacteria are enriched (Reunamo et al. 2013), while in the higher salinity Alpha– and Gammaproteobacteria take over (Viggor et al. 2013).

The bottom sediment may receive a considerable part of the crude oil depending on the vicinity of the contaminating source area and on the sedimentation rate of particles on which the oil compounds are adsorbed (Rousi and Kankaanpää 2012 references therein). After an oil spill the amount of marine snow increases and through the settling different petroleum hydrocarbons enter into higher food webs (Passow et al. 2012). Crude oil is a complex mixture of aliphatic and aromatic hydrocarbon compounds. Aliphatic petroleum hydrocarbons are more readily degraded by the bacterial community than the more complex chemical structures of the polycyclic aromatic hydrocarbons (PAHs) (Bamforth and Singleton 2005). Through sedimentation they both are buried in bottom layers, where the degradation potential of the bacteria may develop slowly and result from the previous exposure to oil–originating compounds (Coates et al. 1997). Petroleum hydrocarbon containing settling matter may lead to death of a part of the sediment bacterial community

2 because of the toxicity of crude–oil compounds (Radniecki et al. 2013) and its transformation products (Bamforth and Singleton 2005).

Anoxic areas are increasing worldwide due to eutrophication and changes in stratification properties of water in the coastal areas (Diaz and Rosenberg 2008). The coastal waters of the Gulf of Finland meet regularly anoxia and bacteria have to switch to alternative electron acceptors or give way to a different bacterial community to take over in petroleum–hydrocarbon degradation. Here, microbial nitrate (Alain et al. 2012) and sulphate (Aeckersberg et al. 1991) reduction processes are responsible for the oxidation of aromatic hydrocarbons in sediments (Coates et al. 1997; Dell'Anno et al. 2009). Also PAHs are degraded in sulphate–reducing conditions. For example, through series of carboxylation and hydrogenation, naphthalene is eventually metabolized to carbon dioxide (reviewed by Bamforth and Singleton 2005). Iron (Fe) reducers may use aromatic hydrocarbons such as toluene and phenol, and Geobacter can degrade cyclic hydrocarbons, which all are difficult for other bacteria to be utilized as energy source in anoxic conditions (Lovley et al. 1989; Zhang et al. 2012). Manganese (Mn) reduction may also be involved in petroleum hydrocarbon degradation (Yeung et al. 2013).

Sediments contain Fe and Mn oxides, but especially high concentrations are found in Fe– Mn concretions, which cover about 10% of the bottom area of the Gulf of Finland (Winterhalter 1966). These mineral precipitates function as a habitat for a diverse bacterial community (Yli–Hemminki et al. 2014), which is exposed to fluctuating redox–conditions in the Gulf of Finland. The concretion microbial community was found to be unique consisting of mainly unknown, uncultured species belonging to 12 different phyla of which the Proteobacteria constituted around 50% (Yli–Hemminki et al. 2014). Additionally, concretions are commonly spherical having a porous structure. Therefore, they may have a considerable surface area for microbial processes. The specific concretion may carry Fe and Mn reducers and the concretion itself supplies the electron acceptors. Crude oil, in turn, may act as an electron donor.

We carried out an experiment (experiment setup 1) to study the degradation potential of concretions for crude oil under oxic and anoxic conditions. Another aim of our study was to compare the degradation potential of concretions to that of surrounding sediment (experiment setup 2). Total PAH degradation and naphthalene mineralization were measured in the experiment 2. In addition, in both of the experiments the abundance of PAH–ring hydroxylating deoxygenase genes (PAH–RHDα) responsible for the initial step of aerobic PAH degradation were measured. Our study gives information about the natural recovery potential of the concretion bottoms in comparison to sediments in varying redox– conditions and the response of the microbial community to the crude–oil contamination.

2. Materials and methods 2.1 Sampling of concretions and sediment Bottom samples for the microcosm experiments were collected from the Gulf of Finland (Table 1) by a van Veen grab sampler and a box corer on board R/V Aranda. Concretions were stored in plastic containers or aerated aquariums filled with near–bottom water. Containers and aquariums were stored at constant 5 °C temperature in darkness. Sediment was stored in aerated aquariums with concretions and near–bottom water.

2.2. Experiment setup 1: Concretions collected from various locations (mixture of concretions) A mixture of spherical concretions of different size, form and origin was weighted (10 g ± 1 WW) in to flasks with 0.01 g (± 0.001–0.004) of Russian crude oil. Ten mL of artificial brackish sea water (described in Yli–Hemminki et al. 2014) was added into the flasks so that concretions were nearly covered with the solution. Crude oil formed a film on the water but was in contact with the concretions. Anoxic flasks were gassed with N2 for five minutes to purge oxygen, whereas oxic flasks were aerated regularly every week. Three replicate oxic (week 0, 2, 4, 6, 7, 11, 15 and 17) and anoxic (week 0, 4, 7 and 17) flasks were emptied by destructive sampling. Samples were taken for total crude–oil analysis, dissolved Fe and Mn measurements and total DNA extraction. The disappearance rate of -1 -1 oil for the mixed–concretions experiment was calculated (μg OIL1 g DWCONC – μg OIL2 g DWCONC)/ (t2 – t1) were “t” is time. There were two types of controls in the experiment: a negative control contained only autoclaved artificial brackish water and crude oil without concretions. There was also an autoclaved control with concretions (autoclaved twice 60 min) for all treatments. Autoclaved and negative controls were aerated weekly. The flasks were incubated in slow agitation (60 rpm) at 10 °C in the darkness to simulate the near– bottom conditions of the coastal Gulf of Finland where temperature varies from 2 °C to 14 °C (Leppäranta and Myrberg 2009).

2.3. Experiment setup 2: Concretions and surrounding sediment from site LL4A Sediment or concretions (10 g WW) or the both (10 g WW total) from site LL4A were weighed into 100–mL flasks in triplicates. Ten milliliters of artificial brackish sea water was added into the flasks. One set of concretions as well as of sediment received nutrition addition (5 mM NH4NO3 and 0.35 mM K2HPO4) twice during the experiment to test the effects of the nutrients on naphthalene mineralization. In the experiment DNA (one set of flasks), and crude–oil and PAH (another set of flasks) analysis were performed only in the beginning (after 1 h settling) and at the end of the experiment (11 weeks). The whole bottle content was emptied by destructive sampling. Samples for dissolved Fe and Mn measurements were taken four times (week 0, 5, 9, 10) from the DNA flasks.

All the experimental sets were incubated at 10 °C, in darkness, and slowly (70 rpm) agitated. In addition, the oxic experiment was aerated once a week by a 60 mL syringe whereas anaerobic experiment was driven anoxic by rinsing the flasks with N2 gas for at least five minutes. Flasks containing sediment and concretions together were autoclaved controls (120 °C for 60 min on three following days).

2.3.1 Mineralization of 14C-naphthalene 14 Mineralization of C–naphthalene to CO2 was studied by radiorespirometry (Tuomi et al. 2004, Nyyssönen et al. 2006). Approximately 3 mg of crude oil and 2 kBq of ring-labelled 14C–naphthalene [specific activity 31.3 mCi/mmol; 99% radiochemical purity (Larodan Fine Chemicals Ab; Sweden)] was added to LL4A concretion or sediment samples in 100– mL infusion bottles. The sealed bottles were incubated for 11 weeks during oxic and 14 anoxic conditions. The production of C–labeled CO2 in the bottles was monitored by 0.5 mL 1 M NaOH traps, which was analysed in a Winspectral 1414 Liquid scintillation counter (Wallac Oy; Turku, Finland) using Ultima Gold high flash point LSC–scintillation cocktail (Perkin Elmer; MA, USA).

2.4 Methods to analyse crude oil, total PAHs and Fe and Mn reduction The samples were analysed for the petroleum hydrocarbon content (sum between C10 – C40) and polycyclic aromatic hydrocarbons (PAH). Petroleum hydrocarbons and PAHs were extracted simultaneously from a mixture of water and concretions or sediment. The extraction of the samples started with addition of 1 mL hexane (Fluka, for pesticide residue analysis) which included n–decane (C10H22, Fluka 30540) and tetracontane (C40H82, Fluka 87086) to check the retention time in the calibration. Then 25 mL hexane was added to the samples which were stirred with a magnetic bar for 30 min. The samples settled about 20 min and the hexane layers were transferred to new tubes. To remove the remaining water from the hexane 0.5 g of Na2SO4 was added. Half of the original extract volume was taken for the PAH analysis and 2 mL was transferred into a gas chromatography vial for the petroleum hydrocarbon analysis. The petroleum hydrocarbon contents were analysed by a gas chromatograph (6890N, Agilent Technologies) equipped with a flame ionization detector (GC–FID) where helium was used as carrier gas. The external calibration solutions were diluted from the mixture made from diesel oil (BAM–K008, DK 1037) and lubrication oil (BAM–K009, HVI 50), both from Shell Global Solutions GmbH, Germany. A gas chromatograph fitted with a HP–5 (length 5 m; i.d. 0.32 mm, film 0.25 μm, Agilent Technologies) column was used. An on–column injection of 1 μL at 60 °Cwas performed and the used oven programme was as follows: 60 °C for 5 min, ramp from 60 °C to 320 °C at 30 °C/ min, ramp from 320 °C to 325 °C at 50 °C/ min and 325 °C for 7 min. Detector temperature was 360 °C.

PAHs were analysed with a gas chromatograph (Varian CP–3800) equipped with a mass spectrometer (Varian 1200). The hexane extract was further concentrated to a volume of 1 mL by gently blowing a stream of nitrogen over the extract. Fractionation of the aliphatic and aromatic compounds was performed with solid phase extraction (SPE) cartridge (Biotage, ISOLUTE EPH, 5 g). The SPE cartridge was conditioned with 20 mL hexane. Of the sample extract 1 mL was added to the SPE cartridge and aliphatic compounds were removed with 12 mL hexane. The PAHs were eluted with 20 mL dichloromethane (J.T. Baker, for organic residue analysis). 1 mL of cyclohexane (Fluka, Pestanal) was added as solvent keeper to the dichloromethane and the solvent was evaporated to 1 mL by gently blowing a stream of nitrogen. The commercial PAH mixture including 18 PAH compounds (Dr.Ehrenstorfer PAH–Mix 9 in cyclohexane) and benzo[e]pyrene and perylene as single compound (both from Dr.Ehrenstorfer in cyclohexane) were used to make calibration solutions to calibrate the instrument. Deuterated PAHs were used as internal standards (Dr.Ehrenstorfer, PAH–Mix 24 in hexane: Naphthalene D8, Acenaphthene D10, Phenanthrene D10, Chrysene D12 and Perylene D12). A gas chromatograph with autosampler (CompiPal, CTC Analytics AG) fitted with a ZB-5MS (length 30 m, i.d. 0.25 mm, film 0.25 μm, Zebron) column and helium as carrier gas was used. An injection (split/splitless injector, 1177, Varian) of 1 μL at 250 °C was performed and the oven programme was as follows: 60 °C for 2 min, ramp from 60 °C to 190 °C at 15 °C/ min, ramp from 190 °C to 250 °C at 4 °C/ min, ramp from 250 °C to 270 °C at 10 °C/ min, ramp from 270 °C to 300 °C at 4 °C/ min, ramp from 300 °C to 320 °C at 20 °C/ min. Selected ion monitoring (SIM) mode with internal calibration was used in the mass spectrometer to identify the compounds.

Liquid aliquots for the analysis of dissolved Fe2+ and total Fe and Mn were taken with a needle– filter (0.45 μm) –syringe–combination through the stoppers of the incubation

5 flasks. Reduced Fe2+ was measured by the ferrozine method with a spectrophotometer at a wavelength of 564 nm (Stookey 1970), and total Fe and Mn were analysed by Inductively Coupled Plasma–Optical Emission Spectrophotometry (ICP–OES).

2.5 DNA extraction and quantification of PAH-degradation genes by qPCR DNA was extracted from 0.5 g of concretions and from the sediment by Fast DNA Spin Kit for Soil (Qbiogene/ MPBiomedicals; Carlsbad, CA, USA), according to the manufacturer’s instructions. The samples were homogenized in FastPrep instrument on setting 4.5 for 30 seconds. The concentration of DNA was measured with Qubit fluorometer (Invitrogen Molecular Probes; Carlsbad, CA, USA). qPCR reactions for the PAH–ring hydroxylating deoxygenase genes (PAH–RHDα) (gram positive and gram negative bacteria separately) were carried out in triplicate (with three technical replicates for each) for the microcosms using a 7300 Real time PCR instrument (Applied Biosystems). The primers used were: gram–positive (GP) forward 5′–CGG CGC CGA CAA YTT YGT NGG–3′ and reverse 5′–GGG GAA CAC GGT GCC RTG DAT RAA–3′ or gram–negative (GN) forward 5′–GAG ATG CAT ACC ACG TKG GTT GGA–3′ and reverse 5′–AGC TGT TGT TCG GGA AGA YWG TGC MGT T–3′ (Cebron et al. 2008), 0.3 μmol each. The 25 μL reaction contained also 12.5 μL Maxima™ SYBR Green qPCR Master Mix (Fermentas, Vilnius, Lithuania), and approximately 7.0–24.0 ng DNA sample. The DNA extracts were diluted 1:10 to reduce the effect of possible PCR inhibition. The values achieved from the qPCR were calculated to dry weight of concretions and sediment. Thermal cycling conditions were 10 min at 95 °C, 42 cycles of 15 s at 95 °C, 30 s at 54 °C, 30 s at 72 °C, and 27 s at 86 °C for GP reaction or 83 °C for GN reaction. After the elongation step a fluorescence reading step was applied in order to exclude signals derived from primer dimer or unspecific products having melting point below 86 °C (GP) or 83 °C (GN). Melting curve analysis was performed after each run. Standards used in the analysis were Escherichia coli plasmids with PAH–RHDα inserts amplified and cloned from Pseudomonas putida G7 (GN) or Mycobacterium vanbaalenii DSM 7251 (GP). Standards with 3 × 10-1 and from 100 to 108 gene copies μL-1 were used for the calibration curve. The detection limit of the assay was 200 gene copies g- 1 sediment or concretion. The amplification efficiency of the qPCR reactions, estimated by using the standard curve, was 100% ±10%.

The average gene copy numbers of PAH–RHDα–genes of gram–positive and gram– negative bacteria were calculated per gram dry weight (DW) of concretions, using 65% dry matter, which was an average of the mixed concretions from different sites and of varying shapes and sizes.

2.6 PCR, cloning and sequencing of the 16S RNA gene The bacterial 16S rRNA gene was amplified with general primers PRBA338f: 5’–AC TCC TAC GGG AGG CAG CAG– ‘3 (Øvreås et al. 1997) and DS907r: 5’ –CCC CGT CAA TTC CTT TGA GTT T– ‘3 (Teske et al. 1996). The PCR reaction (49 μL) constituted of: 31.5 μL mol.–grade water, 5 μL (NH4)2SO4–buffer (10X), 5 μLMgCl2 (25 mM), 1 μL dNTPs (10 mM each), 2 μL primer mix (5 μM), 2 μL Taq polymerase (1 U μL-1) (Fermentas Int., Inc.; Canada), 2.5 μL BSA, and 1 μL of template was used. The amplification was done with PTC–200 DNA Engine (MJ Research; Waltham, MA, USA) in following settings: initial denaturing (3 min at 94 °C) followed by 25 cycles of

6 denaturation (45 s at 94 °C), primer annealing (45 s at 53 °C), and primer extension (1 min at 72 °C). The programme finished with a final extension step (7 min at 72 °C and end at 15 °C). The PCR products were purified using the Wizard SV Gel and PCR Clean–Up System (Promega; Madison, WI, USA) and checked for the right size (~ 569 bp) by gel electrophoresis on a 1.5–% agarose gel (120 V, 60 min).

Prior to cloning, selected triplicate samples of different treatment were compared by DGGE according to Yli–Hemminki et al. (2014) to reveal the variation of the pattern of 16S rDNA between the triplicates. The purified amplification products of pooled triplicate samples were cloned by means of the TOPO TA Cloning Kit for Sequencing, with the plasmid vector of expression PCR 4–TOPO (Invitrogen) according to the manufacturer’s user guide. The cloned inserts were excised for sequencing, with PCR performed directly from bacterial colonies with vector–specific M13 primers. Clones were sequenced at Macrogen (Seoul, South Korea) and chimera check was performed with DECIPHER's Find Chimeras web tool (Wright et al. 2012). The sequences obtained were compared to sequences in the GenBank, and the closest matches were selected. The original sequences were classified on the basis of sequencing results by means of Ribosomal Database Project (RDP) classifier with confidence threshold of 95% (Wang et al. 2007). The sequences determined were submitted to the European Nucleotide Archive and assigned accession numbers from KJ615377 to KJ616364.

2.7 OTU assignment and diversity analysis The Mothur program (Schloss et al. 2009) was used for sequence analysis, Venn diagram analysis, determination the operational taxonomical units (OTUs) and for construction of rarefaction curves. The R statistical software (version 3.1.0) was used for plotting the rarefaction graph. OTUs were assigned at 97% identity threshold and the number of OTUs in the initial concretion and sediment samples was compared by the rarefaction analysis (Fig. S1).

2.8 Statistical analysis We performed one–way ANOVA and Mann–Whitney non-parametric test analysis using 95% confidence interval and pairwise comparisons using 95% confidence interval. The analyses were performed by the SPSS system (version 22, IBM Corp. Released 2013).

3. Results 3.1 Disappearance of crude oil in the experiment 1 In the beginning of the experiment the addition of crude oil to the microcosms resulted in -1 the concentrations of 0.5–0.6 mg (C10–C40 ) g DW with mixed concretions. After seventeen weeks of incubation 59% and 51% of the oil had disappeared in oxic and anoxic conditions, respectively (Table 2). The rate of disappearance was faster in oxic than anoxic conditions during the first two weeks, being 17.6 μg d-1 g-1 DW of concretion (Fig. 1). During anoxia the highest rate of disappearance was 9 μg d-1 g-1 DW. However, the rates stabilized for both treatments after six weeks of incubation (Fig. 1). The concentration of oil diminished 46% also in the flasks containing autoclaved concretions.

3.2 Disappearance of crude oil and PAHs in the experiment 2 In the beginning of the experiment, the addition of crude oil to the microcosms resulted in -1 the concentrations of 0.2–0.4 mg (C10–C40) g DW with concretions or surrounding

7 sediment from the site LL4A. During the oxic conditions, around 70% of the added crude oil disappeared from the flasks containing LL4A concretions (eleven weeks incubation, Table 2). In the corresponding sediment flasks the amount of crude oil decreased 54–80% within eleven weeks. During anoxic conditions, the disappearance values 41–47% for sediment were in the range of 35–61% obtained for the concretions (Table 2). Crude oil diminished also in the autoclaved controls, including both LL4A sediment and concretions, by 10% in the anoxic and 80% in the oxic experiment.

The concentration of 18 different PAH compounds was analysed in the LL4A experiment (Table S1 and S2). The analysis included 1–methylnaphthalene and 2–methylnaphthalene, which are not among the 16 commonly analysed EPA PAHs. These two compounds constituted >95% of the total PAHs in the crude oil in this study. The concentrations of PAHs diminished (96%) in all incubations containing concretions or sediment, both in oxic and anoxic conditions. Even a proportion of high molecular weight PAHs diminished during the experiment. The disappearance of PAHs was up to 79% also in the autoclaved oxic controls. However, the disappearance pattern of PAHs differed between the autoclaved and the non–autoclaved microcosms (Fig. 2). In the non–autoclaved microcosms mostly 4–, 5–, and 6–ringed PAHs remained, because the 2– and 3 –ringed PAHs were degraded. Whereas in the autoclaved controls the 2–,3–, and 4–ringed PAHs still dominated in the end of the experiment (Fig. 2).The PAH disappearance results were similar in the anoxic conditions, except that only 35% of PAHs disappeared in the autoclaved controls (Table 2).

3.2.1 Mineralization of 14C-naphthalene Mineralization of 14C–naphthalene was measured from the LL4A incubations (experiment 2). In the oxic concretion flasks approximately 5–8% of the 14C–naphthalene was mineralized (Table 2, Fig. 3a).The addition of nutrients did not enhance the mineralization. In the anoxic microcosms around 2–4% of the 14C–naphthalene was mineralized (Table 2, Fig. 3b).

Around 5–14% of the 14C–naphthalene was mineralized by the oxic sediment (Table 2, Fig. 3a). The nutrient addition did not enhance the mineralization in the sediments. In anoxic conditions the mineralization remained very low (< 3%) (Table 2, Fig. 3b). The mineralization of 14C–naphthalene started rapidly in the control microcosms without crude oil addition.

In general, the mineralization of 14C–naphthalene was significantly (p < 0.05) higher in the oxic than anoxic conditions. Only 0.2% 14C–naphthalene mineralization was observed in the autoclaved microcosms (Table 2, Fig. 3).

3.3 Quantification of PAH-degradation genes 3.3.1 Experiment 1 In Experiment 1 with mixed concretions, the copy number of gram positive PAH ring– hydroxylating dioxygenase (GP PAH–RHDα) increased in the both oxic (1.8×105 ±9.4×104) and anoxic (1.3×105 ±1.2×105) incubations during the first four weeks, but there was no significant difference. In the end of the experiment, the GP PAH–RHDα and the gram–negative (GN) gene copy levels were approximately at the same level as in the initial samples (Table 2).

3.3.2 Experiment 2 In the oxic LL4A concretion microcosms, the number of GN PAH–RHDα gene copies increased 1000–fold and the gene copy number was significantly higher (p < 0.05) than in the initial concretion sample (Table 2). Additionally, the number of GN PAH–RHDα gene copies was higher in the concretion than in the sediment flasks (p < 0.05). In the anoxic experiment, the number of GN PAH–RHDα gene copies was significantly higher than in the initial concretion sample (p < 0.05), but no increase was detected in the GP gene copy numbers (Table 2). Nutrient addition did not have significant effect on the number of GN or GP PAH–RHDα gene copies in the concretion flasks.

In the oxic LL4A sediment experiment the number of GN PAH–RHDα gene copies increased significantly (p < 0.05) in the flasks with nutrients, when compared to the initial sediment (Table 2).

3.4 Dissolution of iron and manganese from the concretions and sediment Fe and Mn dissolved with great variation between three replicates containing mixed concretions. Fe2+ constituted 90% of the total dissolved Fe accumulated in the solution. Concentrations of dissolved Fe ranged between 93–800 μg L-1 in anoxic treatments indicating that efficient Fe reduction had started only in few anoxic flasks (one replicate out of three, after seven and seventeen weeks). On the contrary, the concentrations of dissolved Mn were up to hundred times higher than Fe, varying between 1–74 mg L-1.

Generally the concentrations of dissolved Fe remained below 0.2 mg L-1 for both the LL4A concretions and sediment in anoxic and oxic conditions. The concentrations of dissolved Mn were higher, but stayed below 3 mg L-1. Dissolution of both Fe and Mn was higher in the autoclaved controls.

3.5 Characterisation of the bacterial communities Selected bacterial communities were first compared by DGGE. Cluster analysis of DGGE bands showed that the bands varied less between replicates than the treatments (data not shown). Therefore, the triplicate DNAs were pooled before preparing the clone library. Altogether, 11 clone libraries were constructed and 987 sequences obtained. The cloned sequences were classified by means of RDP classifier with confidence threshold of 95% for classification to bacteria.

The changes in the microbial community of concretions and sediment were observed prior to and after the addition of petroleum hydrocarbons. Half of the initial concretion (both LL4A and mixed) bacterial sequences and 42–54% of sediment sequences were affiliated to Proteobacteria (Fig.4). About 35–40% of concretion bacterial sequences and 20–29% of sediment bacterial sequences remained unclassified. Initial concretion and sediment samples were also different in the proportions of other bacterial phyla. Actinobacteria, Bacteroidetes and Chloroflexi were more abundant in sediment than in concretions.

The distribution of Proteobacteria classes differed between the concretion and sediment samples (Fig.5). Gammaproteobacteria were initially dominant in the concretions (30– 38%) whereas Deltaproteobacteria were dominant (30–60%) in the sediment. Epsilonproteobacteria occurred only in sediment samples (4% of the Proteobacteria). The

9 proportion of unclassified Proteobacteria was highest in the mixed concretion samples (32% of the Proteobacteria).

The proportions of unclassified bacteria decreased 10% or more in all experiments with added crude oil (Fig. 4). The proportions of Proteobacteria increased in all treatments 8– 30% (Fig. 4). Distinct distribution patterns for concretions and sediment were observed also after incubations. The proportion of Nitrospira increased 8% in one concretion treatment (Fig. 4). Actinobacteria, Chloroflexi and Deltaproteobacteria declined from sediment treatments (Fig. 4). In concretions mainly the proportions of Alpha– and Betaproteobacteria were increased and in sediment mainly of Beta– and Gammaproteobacteria (Fig. 5).

A total number of 515 OTUs were obtained in the 11 clone libraries sequenced (at 0.03 distance). Around 8.9% OTUs were shared between the initial LL4A concretions and sediment, but only 2.3% between the initial samples of mixed concretions and LL4A sediment. Rarefaction curves (Fig. S1) revealed that the number of OTUs were somewhat higher in the initial sediment than in the initial concretions. Compared to the initial samples the number of OTUs remained approximately the same after different treatments.

The major fraction (99%) of the cloned sequences matched to uncultured clones which seldom could be classified to bacterial order or family. The cloned sequences were compared to their closest matches in the GenBank, and categorized into ecotypes based on the environment they were obtained from (Fig. 6).

The pattern of matching original environments was similar for all concretion samples, despite of their origin or experimental treatments (Fig. 6). A large proportion of the clones from the initial concretions was similar (90–100% similarity) to the sequences earlier found from the Baltic Sea concretions (Yli–Hemminki et al. 2014). These clones could be classified as Gamma-, Beta- and Deltaproteobacteria (Fig. 6, Table 3). In general, after oxic incubation of concretions, the distribution of clones between the different environments of origin shifted slightly (Fig.6), and the closest matches were different (Table S3). In the anoxic conditions (mixed concretions) clones classified as Albidoferax (Comamonadaceae, Betaproteobacteria) enriched and their closest matches were found to uncultured cloned sequences from a pyrite mine (Table 3). The clones classified as Nitrospira (Fig. 4) matched to sequences originating from different environments: Antarctic, As–contaminated sediment, Fe– and S–precipitating microbial mat (Omoregie et al. 2008) and lava.

Sediment samples showed a different pattern. Only few clones from the sediment experiments matched to the sequences from the Fe–Mn concretions of the Baltic Sea (Yli– Hemminki et al. 2014, Fig. 6, Table 3). After oxic incubation of sediment the proportion of clones matching to sequences originating from sediment and oil–contaminated sites, as well as from water or waste water origin, increased. Clones classified as Gammaproteobacteria, similar to sequences originating from oil–contaminated sediment (Paisse et al. 2010) and Baltic Sea sediment enriched (Edlund et al. 2008). After anoxic incubation of sediment there was also increase in clones which matched to sequences originating from metal–containing environments such as Mn–rich sediment (Vandieken et al. 2012). Three clones were affiliated with cultured species Geopsychrobacter

10 electrodiphilus (Deltaproteobacteria) (Holmes et al. 2004) known to reduce iron and to transfer electrons to solid stage through pili.

4. Discussion Concretions and sediment are physically and chemically different environments for microbes; concretions being solid but porous metal precipitates, whereas sediment comprised mainly of fine grained clay containing metal oxides. Rather than to study oil– degrading bacterial strains in laboratory conditions at high temperatures, the aim of the study was to investigate the overall degradation capacity in different bottom types at temperature which occur in the coastal Gulf of Finland.

Degradation of crude oil Hydrocarbon–degrading bacteria were present in the both concretions and sediment. Fairly similar degradation potentials for crude oil between concretions and sediment in both oxic and anoxic conditions were observed. Oil degradation started within a week and more than 50% of the added crude oil disappeared within three months in the experiments at 10 °C.

Disappearance of the added crude oil from the autoclaved microcosms may partly be explained by the fact that autoclaving changed the chemical structure of the concretions and sediment and thus altered their reactivity or sorption properties (Tuominen et al 1994). Thus, autoclaving may have caused release of Fe and Mn into the solution and may therefore not be a sufficient method to form control incubations for abiotic reactions. Iron–bearing minerals are reactive and can oxidise alkanes and other hydrocarbons under certain conditions (Surdam et al. 1993; Seewald 2003; Ferrarese et al .2008). This is supported by the stable concentrations of oil in the aerated control incubations containing only water and crude oil (Table 2). In these units the concentration remained around 6 mg per incubation unit during the experiment. The oil disappearance rates calculated for the mixed concretion microcosms were fast (Fig. 1), whereas in the autoclaved controls the decrease in the crude–oil amount happened drastically once after seven weeks and then stayed at a constant level. Even if long incubation time may cause recovery of bacteria in the autoclaved samples, and the added crude oil may have contained bacteria, no signs of the recovery were observed in naphthalene mineralization or in the disappearance pattern of total PAH fractions.

Considering the concentrations of the treated and autoclaved control units, the proportion of biological degradation in Experiment 1 may have been at least 13% and 5% in oxic and anoxic incubations, respectively, if a conservative estimate of the abiotic disappearance is invoked. However, the abiotic controls may have been altered and this overestimates the abiotic disappearance.

In the autoclaved controls of Experiment 2 (LL4A concretions with sediment) about 10% (anoxic) and 80% (oxic) of crude oil diminished in the experiment. Based on these values the proportion of biological degradation in anoxic conditions may have been at least around 25–50% with the concretions and around 30% with the sediment.

Nutrient addition increased the disappearance of crude oil only slightly in contrast to other studies where fertilization has been shown to enhance biodegradation significantly (Coulon et al. 2007; Pelletier et al. 2004; Lee and Levy 1989). In addition, fertilization has been

11 shown to have more important effect on biodegradation and bacterial community than temperature (Rodriguez–Blanco et al. 2010). Temperature in the bottom of the coastal Gulf of Finland varies between 4–10 °C, which can slow down the degradation of complex compounds. Also, it has been shown that higher nutrient additions enhance the oil degradation in the beginning of experiments, but at the end, the same amount is degraded even in experimental units with low nutrient additions (Röling et al. 2002). Degradation is not likely to be limited by nutrients in the Gulf of Finland, because the concentrations of N and P in the surface sediment can be 12 mg and 4 mg g-1 DW, respectively (Lehtoranta 1998). Concretions contain high amounts of P (28 mg g-1 DW, Ingri 1985) which may dissolve and become available for oil–degrading bacteria. Therefore, it is possible that the nutrients were released from the concretions and sediments and became readily bioavailable for oil–degrading bacteria.

Degradation of PAHs and mineralization of naphthalene Naphthalene, and even more complex PAH compounds, were removed efficiently by concretions and sediment in oxic and anoxic treatments (Table 2). The ring–hydroxylating dioxygenase (RHDα) involved with the initial step of PAH metabolism is widespread in bacteria, and is able to move among bacterial genera by horizontal gene transfer (Hickey et al. 2012). The common existence of RHDα may explain part of the removal of naphthalene and PAHs, but the direct chemical oxidation or evaporation of PAHs to the airspace of the flasks (Leduc et al. 1992) may have also contributed to the disappearance. It has been evaluated that volatilization can account approximately 30% and abiotic reactions 2% loss of naphthalene in soil systems (Park et al. 1990). However, in our experiment, the disappearance patterns of total PAHs fractions were different between non–sterile and autoclaved controls, which indicated that mineralisation had an important role in the disappearance of the PAHs. Based on the differences in decrease, the proportion of biological PAH degradation in oxic conditions would then be at least17% and the remaining part might be explained with evaporation or other non–biological effect in the autoclaved controls. The disappearance of PAHs was the lowest in the autoclaved incubations, and especially in anoxic units. The percentage of biological PAH degradation in anoxic conditions is thus considered to be at least 60%, and the rest would be due to evaporation or other non–biological effect in the autoclaved controls. Additionally, aromatic compounds are not very sensitive towards abiotic removal. When testing a wide variety of aromatic compounds Lovley et al. (1989) have found out that only catechol reduced Fe3+ abiotically at near–neutral pH values. Abiotic degradation of naphthalene has been studied also in the presence of Mn oxides in soil–water system, but no degradation was detected after nine weeks incubation (Mihelcic and Luthy 1988).

The abundance of the PAH–degradation genes in the microcosms The gene copy number of PAH–RHDα is a measure of the degradation potential of the bacterial community under the pressure of aromatic compounds in crude oil (Cebron et al. 2008), although it does not reveal the actual activity (transcription) of the gene. Traces of oxygen in the beginning of the experiment may have increased the gene copy numbers also in anoxic conditions. It is beneficial for bacteria to keep the gene also in the anoxic conditions, as the oxygen levels in the sediment may fluctuate. Also, Tuomi et al. (2004) have observed the presence of aerobic PAH–degradation genes in anoxic petroleum– hydrocarbon contaminated subsurface soil. In Experiment 2 concretions the number of GN PAH–RHDα gene increased four orders of magnitude in the oxic experiment. In

Experiment 1 significant increase was not observed. The higher concentration of crude oil may have caused toxic effects (Rahman et al. 2002) in these incubations and bacteria may have favoured more easily degradable crude–oil compounds over PAHs. The GP PAH– RHDα gene copy numbers remained approximately at the initial levels in all flasks, increasing only temporarily ten–fold in the oxic incubation of mixed concretions. The notable increase in GN PAH–RHDα gene copies is in line with the observed mineralization of PAHs, and with the increasing large proportion of gram negative Proteobacteria in the oil exposed samples.

Fe and Mn dissolution from concretions The addition of crude oil did not lead to efficient dissolution of Fe and Mn in Experiment 2. Nitrate was added as nutrient only in Experiment 2, and it may have acted as electron acceptor taking over Fe and Mn in these incubations (LaRiviere et al. 2001; Wang et al. 2012). It is possible that metal–reducing bacteria need more time to adapt to oil contamination in sediments (Coates et al. 1997, Anderson and Lovley 1999). Therefore, it may be more relevant to keep bacteria in contact with crude oil than to focus on serving Mn4+ (Li et al. 2011) or Fe3+ (Li et al. 2010) as electron acceptors for the bacteria at a polluted site. At high concentrations (e.g., 10 mg Mn L-1), metals may even be toxic to sediment bacteria (Li et al. 2011) if they have not been adapted to such concentrations. Different micro niches may have developed inside and on the surface of concretions as well as in sediment layers. The remaining oxygen was consumed in the sealed flasks and bacteria have had to switch to use alternative electron acceptors in anoxic conditions. The high concentrations of dissolved Mn from mixed concretions indicated that especially Mn could have acted as terminal electron acceptor. Mn was shown to be used for respiration in cold–climate groundwater system contaminated by petroleum hydrocarbons (Yeung et al. 2013) and in soil percolation columns (Langenhoff et al. 1996). In few anoxic incubation flasks of mixed concretions, Fe2+ could have dissolved due to anaerobic bacterial reduction, and Fe reduction have been shown to be coupled to anaerobic mineralization of aromatic contaminants (Lovley et al. 1989; Lovley et al. 1994), such as naphthalene in sediments under Fe3+–reducing conditions (Anderson and Lovley 1999). Sulphate reduction may have been insignificant, because there was no smell of H2S or formation of black precipitates indicating presence of Fe sulphides. In the Experiment 2 autoclaving may have caused the great dissolution of Fe and Mn measured from LL4A concretions, and sediments as higher dissolution was observed in the autoclaved controls.

Bacterial community structures in concretion and sediment microcosms Different bacterial communities enriched in concretions and surrounding sediment during the incubations, but their ability to degrade crude oil was similar. The crude oil is a mixture of different compounds and therefore different bacteria may have been needed to degrade them in concert (Dell’Anno et al. 2012). The overall number of different bacterial phyla of concretions and sediment was high even after several weeks of oil exposure; and compared with the initial samples the number of OTUs stayed approximately the same after different treatments. The development of different bacterial communities in concretions and sediments indicates that various communities are capable to tolerate and degrade the compounds of crude oil. In many previous studies (Röling et al. 2002; McKew et al. 2007; reviewed in Head et al. 2006 and Yakimov et al. 2007) bacterial diversity has been shown to decrease during oil exposure favouring the enrichment of known oil– degrading species such as Thalassolituus oleivorans, Cycloclasticus and Alcanivorax.

However, in a more recent study, diverse bacterial communities in the sediment have been observed to correlate positively with the hydrocarbon biodegradation efficiency (Dell'Anno et al. 2012).

The number of common OTUs was less than 10% between concretions and sediments. The rising slope of the rarefaction curves (Fig. S1) indicated that the number of OTUs would still increase if more sequences were cloned. In general, the closest matches of the sequences obtained from the concretions were found from various environments and there was no clear shift towards known oil–degrading species. Ecotypes rather than species (OTUs) enriched in all the experiments which was dependent on metal–rich concretions and surrounding sediments. The results of this more extensive study confirm the earlier findings that the Baltic Sea concretions maintain a distinct community with a high diversity and many still unknown species (Table 3), which are only distantly related to any cultured species. The clear high frequency of clones affiliated with specific α– β– γ– and δ–proteobacteria may indicate that we have new clades of bacteria within the Rhizobiales and Rhodospirillales of the α–proteobacteria, Burkholderiales–incetertae–sedis of the β– proteobacteria, Cromatiales of the γ–proteobacteria, and the Syntrophobacteriales of the δ –proteobacteria.

Crude–oil treatment of sediments led to the shift of clones towards sequences (classified as Gammaproteobacteria) previously obtained from oil–contaminated environments (Fig. 6, Table S3). In contrast, the addition of crude oil to concretions increased the abundance of uncultured bacterial clones found from the concretions themselves and metal–rich sites. Some of these clones and the sites where they appear have nevertheless been linked to oil contamination previously. Sequences classified as Alphaproteobacteria from oxic incubation matched sequences originating from polluted soil (Table 3) and Mn–oxide minerals (Miller et al. 2012). For example Betaproteobacteria obtained from anoxic incubations of mixed concretions could be classified as Comamonadaceae, a family which includes the known Fe–reducing bacteria classified as Albidiferax (Ramana and Sasikala 2009). The result indicates that concretions of the Baltic Sea are a characteristic habitat which may form a specific community of microbes different from the surrounding sediment, also capable to remove oil.

Conclusions The concretion bottoms in the Baltic Sea have a substantial potential to remove crude oil and naphthalene at 10 °C temperature similar to the surrounding sediment. The disappearance rate of crude oil was faster during oxic than anoxic conditions but both the concretions and sediment were able to remove half of the added crude oil in oxic and anoxic conditions. There was evidence on the significant biodegradation for both concretions and sediments. Crude–oil addition led to development of different bacterial communities in concretions compared to those of sediment. The numbers of OTUs did not decrease prior to crude oil addition. This indicates that both the concretions and sediment are able to maintain rich bacterial communities typical to these environments, having potential to degrade crude oil.

Acknowledgements This work was financially supported by K.H. Renlund Foundation, Finnish Cultural Foundation, Maa– ja vesitekniikan tuki ry and MUTKU ry. The authors wish to thank Suvi

Tyynelä for her contribution to the experimental work. We extend our gratitude to Anne Markkanen for performing the oil analysis and Timo Sara–Aho for metal analytics (Laboratory, SYKE).

References Aeckersberg F, Bak F, Widdel F.1991. Anaerobic oxidation of saturated hydrocarbons to CO2 by a new type of sulfate–reducing bacterium. Arch Microbiol 156:5–14. Alain K, Harder J, Widdel F, Zengler K. 2012. Anaerobic utilization of toluene by marine alpha– and gammaproteobacteria reducing nitrate. Microbiology 158:2946–2957. Anderson RT, Lovley DR. 1999. Naphthalene and benzene degradation under FeIII– reducing conditions in petroleum–contaminated aquifers. Bioremediation J 3:121– 135. Bamforth SM, Singleton I. 2005. Bioremediation of polycyclic aromatic hydrocarbons: current knowledge and future directions. J Chem Technol Biot 80:723–736. Brakstad OG, Lodeng AGG. 2005. Microbial diversity during biodegradation of crude oil in seawater from the North Sea. Microb Ecol 49:94–103. Cappello S, Denaro R, Genovese M, Giuliano L, Yakimov MM. 2007. Predominant growth of Alcanivorax during experiments on "oil spill bioremediation" in mesocosms. Microbiol Res 162:185–190. Cebron A, Norini MP, Beguiristain T, Leyval C. 2008. Real–Time PCR quantification of PAH–ring hydroxylating dioxygenase PAH–RHD alpha genes from Gram positive and Gram negative bacteria in soil and sediment samples. J Microbiol Meth 73:148–159. Coates JD, Woodward J, Allen J, Philp P, Lovley DR. 1997. Anaerobic degradation of polycyclic aromatic hydrocarbons and alkanes in petroleum–contaminated marine harbor sediments. Appl Environ Microbiol 63:3589–3593. Coulon F, McKew BA, Osborn AM, McGenity TJ, Timmis KN. 2007. Effects of temperature and biostimulation on oil–degrading microbial communities in temperate estuarine waters. Environ Microbiol 91:177–186. Dell'Anno A, Beolchini F, Gabellini M, Rocchetti L, Pusceddu A, Danovaro R. 2009. Bioremediation of petroleum hydrocarbons in anoxic marine sediments: Consequences on the speciation of heavy metals. Mar Pollut Bull 59:1808–1814. Dell'Anno A, Beolchini F, Rocchetti L, Luna GM, Danovaro R. 2012. High bacterial biodiversity increases degradation performance of hydrocarbons during bioremediation of contaminated harbor marine sediments. Environ Pollut 167:85– 92. Diaz RJ, Rosenberg R. 2008. Spreading dead zones and consequences for marine ecosystems. Science 321:5891, 926–929. Edlund A, Hardeman F, Jansson JK, Sjoling S. 2008. Active bacterial community structure along vertical redox gradients in Baltic Sea sediment. Environ Microbiol 10:2051–2063. Ferrarese E, Andreottola G, Oprea IA. 2008. Remediation of PAH–contaminated sediments by chemical oxidation. J Haz Mat 152:128–139. Gutierrez T, Singleton DR, Berry D, Yang T, Aitken MD, Teske A. 2013. Hydrocarbon– degrading bacteria enriched by the Deepwater Horizon oil spill identified by cultivation and DNA–SIP. ISME J 711:2091–2104.

Hao C, Zhang L, Wang L, Li S, Dong H. 2012. Microbial community composition in acid mine drainage lake of Xiang mountain sulfide mine in Anhui province, China. Geomicrobiol J 29:886–895. Hazen TC, Dubinsky EA, DeSantis TZ, Andersen GL, Piceno YM, Singh N, Jansson JK et al. 2010. Deep–Sea oil plume enriches indigenous oil–degrading bacteria. Science 330:204–208. Head IM, Jones DM, Roling WFM. 2006. Marine microorganisms make a meal of oil. Nat Rev Microbiol 43:173–182. Helcom 2012. Clean Seas Guide, The Baltic Sea Area, MARPOL 73/78 Special Area. In Stankiewicz M. ed.. Katajanokankaari 6 B FI–00160 Helsinki, Finland. Hickey WJ, Chen S, Zhao J. 2012. The phn island: a new genomic island encoding catabolism of polynuclear aromatic hydrocarbons. Frontiers in Microbiology 3:1– 15. Holmes DE, Nevin KP, Lovley DR. 2013. Comparison of 16S rRNA, nifD, recA, gyrB, rpoB and fusA genes within the family Geobacteraceae fam. nov Int J Syst Evol Microbiol 54 PT 5:1591–1599. Ingri J. 1985. Geochemistry of ferromanganese concretions and associated sediments in the Gulf of Bothnia. Doctoral thesis, Luleå University of Technology, Sweden. Langenhoff AAM, Zehnder AJB, Schraa G.1996. Behaviour of toluene, benzene and naphthalene under anaerobic conditions in sediment columns. Biodegradation 7:267–274. LaRiviere D, Harris BC, Autenrieth RL, Bonner JS, Keith N. 2001. Electron acceptor consumption during petroleum degradation in microcosm experiments. In Phytoremediation, wetlands, and sediments: The sixth in situ and on–site bioremediation symposium. Leeson, A., Foote, E.A., Banks, M.K. and Magar, V.S. editors. 25–32. Battelle Press, U.S.A. ISBN 1–57477–115–9. Leahy JG, Colwell RR. 1990. Microbial degradation of hydrocarbons in the environment. Microbiol Rev 543:305–315. Lee K, Levy EM. 1989. Enhancement of the natural biodegradation of condensate and crude oil on beaches of Atlantic Canada. International Oil Spill Conference Proceedings: February 1989, 1:479–486. Leduc R, Samson R, Al–Bashir B, Al–Hawari J, Cseh T. 1992. Biotic and abiotic disappearance of four PAH compounds from flooded soil under various redox conditions. Water Science Technol 26:51–60. Lehtoranta J. 1998. Net sedimentation and sediment–water nutrient fluxes in the eastern Gulf of Finland, Baltic Sea. Vie et Milieu 48:341–352. Leppäranta M, Myrberg K. 2009. Physical oceanography of the Baltic Sea. Springer– Praxis books in geophysical sciences. ISBN 978–3–540–79702–9. Li C–H, Wong Y–S, Tam NF–Y. 2010. Anaerobic biodegradation of polycyclic aromatic hydrocarbons with amendment of ironIII in mangrove sediment slurry. Bioresource Technol 101:8083–8092. Li C–H, Wong Y–S, Tam NF–Y. 2011. Effect of MnIV on the biodegradation of polycyclic aromatic hydrocarbons under low–oxygen condition in mangrove sediment slurry. J Hazard Mater 190:786–793. Lovley DR, Baedecker MJ, Lonergan DJ, Cozzarelli IM, Phillips EJP, Siegel DI. 1989. Oxidation of aromatic contaminants coupled to microbial iron reduction. Nature 339:297–300.

Lovley DR, Woodward JC, Chapelle FH. 1994. Stimulated anoxic biodegradation of aromatic hydrocarbons using FeIII ligands. Nature 370:128–131. McGenity TJ. 2014. Hydrocarbon biodegradation in intertidal wetland sediments. Curr Opin Biotech 27:46–54. McKew BA, Coulon F, Osborn AM, Timmis KN, McGenity TJ. 2007. Determining the identity and roles of oil–metabolizing marine bacteria from the Thames estuary, UK. Environ Microbiol 91:165–176. Mihelcic JR, Luthy RG. 1988. Degradation of polycyclic aromatic hydrocarbon compounds under various redox conditions in soil–water systems. Appl Environ Microbiol 545:1182–1187. Miller AZ, Dionísio A, Sequeira Braga MA, Hernández–Mariné M, Afonso MJ, Muralha VSF et al. 2012. Biogenic Mn oxide minerals coating in a subsurface granite environment. Chem Geol 322–323:181–191. Nielsen MB, Kjeldsen KU, Lever MA, Ingvorsen K. 2014. Survival of prokaryotes in a polluted waste dump during remediation by alkaline hydrolysis. Ecotoxicology 233:404–418. Nyyssönen M, Piskonen R, Itävaara M. 2006. A targeted real–time PCR assay for studying naphthalene degradation in the environment. Microb Ecol 523:533–543. Omoregie EO, MastalerzV, de Lange G, Straub KL, Kappler A, Roy H, Stadnitskaia A, Foucher JP, Boetius A. 2008. Biogeochemistry and community composition of iron– and sulfur–precipitating microbial mats at the Chefren mud volcano Nile Deep Sea Fan, Eastern Mediterranean. Appl Environ Microb 74:3198–3215. Paisse S, Goni–Urriza M, Coulon F, Duran R. 2010. How a bacterial community originating from a contaminated coastal sediment responds to an oil input. Microb Ecol 60:394–405. Park KS, Sims RC, Dupont RR, Doucette WJ, Matthews JE. 1990. Fate of PAH compounds in two soil types: Influence of volatilization, abiotic loss and biological activity. Environ Toxicol Chem 9:187–195. Passow U, Ziervogel K, Asper V, Diercks A. 2012. Marine snow formation in the aftermath of the Deepwater Horizon oil spill in the Gulf of Mexico. Environ Res Lett 7. Pelletier E, Delille D, Delille B. 2004. Crude oil bioremediation in sub–Antarctic intertidal sediments: chemistry and toxicity of oiled residues. Mar Environ Res 57:311–327. Radniecki TS, Schneider MC, SempriniL. 2013. The influence of Corexit 9500A and weathering on Alaska North Slope crude oil toxicity to the ammonia oxidizing bacterium, Nitrosomonas europaea. Mar Pollut Bull 68:64–70. Rahman KSM, Thahira–Rahman J, Lakshmanaperumalsamy P, Banat IM. 2002. Towards efficient crude oil degradation by a mixed bacterial consortium. Bioresource Technol 85:257–261. Ramana CV, Sasikala C. 2009. Albidoferax, a new genus of Comamonadaceae and reclassification of Rhodoferax ferrireducens Finneran et al., 2003 as Albidoferax ferrireducens comb. nov. J Gen Appl Microbiol 55:301–4. Reunamo A, Riemann L, Leskinen P, Jørgensen KS. 2013. Dominant petroleum hydrocarbon–degrading bacteria in the Archipelago Sea in South–West Finland Baltic Sea belong to different taxonomic groups than hydrocarbon degraders in the oceans. Mar Pollut Bull 72:174–180. Rodriguez–Blanco A, AntoineV, Pelletier E, Delille D, Ghiglione F. 2010. Effects of temperature and fertilization on total vs. active bacterial communities exposed to

crude and diesel oil pollution in NW Mediterranean Sea. Environ Pollut 158:663– 673. Rocker D, Brinkhoff T, Gruner N, Dogs M, Simon M. 2012. Composition of humic acid– degrading estuarine and marine bacterial communities. FEMS Microb Ecol 80:45– 63. Rotaru C, Woodard TL, Choi S, Nevin KP. 2012. Spatial heterogeneity of bacterial communities in sediments from an infiltration basin receiving highway runoff. Microb Ecol 64:461–473. Rousi H, Kankaanpää H. Eds. 2012. The ecological effects of oil spills in the Baltic Sea – the national action plan of Finland. Environmental Administration Guidelines 6/2012, Nature, 88 p. Finnish Environment Institute. URN:ISBN 978–952–11– 4129–4, ISBN 978–952–11–4129–4 PDF. http://hdl.handle.net/10138/41546 Röling WFM, Milner MG, Jones DM, Lee K, Daniel F, Swannell RJP, Head IM. 2002. Robust hydrocarbon degradation and dynamics of bacterial communities during nutrient–enhanced oil spill bioremediation. Appl Environ Microbiol 68:5537– 5548. Schloss PD, Westcott SL, Ryabin T, Hall JR, Hartmann M, Hollister EB et al. 2009. Introducing mothur: Open–source, platform–independent, community–supported software for describing and comparing microbial communities. Appl Environ Microbiol 75:7537–41. Seewald JS. 2003. Organic–inorganic interactions in petroleum–producing sedimentary basins. Nature 426:327–333. Stookey LL. 1970. Ferrozine – a new spectrophotometric reagent for iron. Anal Chem 42:779–781. Surdam RC, Jiao ZS, MacGowan DB. 1993. Redox reactions involving hydrocarbons and mineral oxidants: A mechanism for significant porosity enhancement in sandstones. Bull Am Assoc Petrol Geol 77:1509–1518. Teeling, H., Fuchs, B.M., Becher, D., Klockow, C., Gardebrecht, A., Bennke, C.M. et al., 2012. Substrate–controlled succession of marine bacterioplankton populations induced by a phytoplankton bloom. Science, 336, 608–611. Teske A, Wawer C, Muyzer G, Ramsing NB. 1996. Distribution of sulfate–reducing bacteria in a stratified fjord Mariager fjord, Denmark as evaluated by most– probable–number counts and denaturing gradient gel electrophoresis of PCR– amplified ribosomal DNA fragments. Appl Environ Microbiol 62:1405–1415. Tuomi PM, Salminen JM, Jørgensen KS. 2004. The abundance of nahAc genes correlates with the 14C–naphthalene mineralization potential in petroleum hydrocarbon– contaminated oxic soil layers. FEMS Microb Ecol 51:99–107. Tuominen L, Kairesalo T, Hartikainen H. 1994. Comparison of methods for inhibiting bacterial activity in sediment. Appl Environ Microbiol 60:3454-3457. VandiekenV, Pester M, Finke N, Hyun JH, Friedrich MW, Loy A, Thamdrup B. 2012. Three manganese oxide–rich marine sediments harbor similar communities of acetate–oxidizing manganese–reducing bacteria. ISME J 6:2078–2090. Viggor S, Juhanson J, Joesaar M, Mitt M, Truu J, Vedler E, Heinaru A. 2013. Dynamic changes in the structure of microbial communities in Baltic Sea coastal seawater microcosms modified by crude oil, shale oil or diesel fuel. Microbiol Res 168:415–427.

Wang Y, Wan R, Zhang SY, Xie SG. 2012. Anthracene biodegradation under nitrate– reducing condition and associated microbial community changes. Biotechnol Bioproc E 17:371–376. Wang Q, Garrity GM, Tiedje JM, Cole JR. 2007. Naïve Bayesian classifier for rapid assignment of rRNA sequences into the new bacterial taxonomy. Appl Environ Microbiol 73:5261–7. Winterhalter B. 1966. Pohjanlahden ja Suomenlahden rautamangaanisaostumista. Geoteknillisiä julkaisuja No. 69, Geologian tutkimuslaitos, Otaniemi, Finland. Wright ES., Yilmaz LS, Noguera DR. 2012. "DECIPHER, A Search-Based Approach to Chimera Identification for 16S rRNA Sequences." Applied and Environmental Microbiology, doi:10.1128/AEM.06516-11. Wu Y–H, Liao L, Wang C–S, Ma W–L, Meng F–X, Wu M, Xu X–W. 2013. A comparison of microbial communities in deep–sea polymetallic nodules and the surrounding sediments in the Pacific Ocean. Deep–Sea Res I 79:40–49. Yakimov MM, Timmis KN, Golyshin PN. 2007. Obligate oil–degrading marine bacteria. Curr Opin Biotech 18:257–266. Yeung CV, Van Stempvoort DR, Spoelstra J, Bickerton G, Voralek J, Greer CW. 2013. Bacterial community evidence for anaerobic degradation of petroleum hydrocarbons in cold climate groundwater. Cold Reg Sci Technol 86:55–68. Yli–Hemminki P, Jørgensen KS, Lehtoranta J. 2014. Iron–manganese concretions sustaining microbial life in the Baltic Sea: The structure of the bacterial community and enrichments in metal–oxidizing conditions. Geomicrobiology J 31:263–271. Øvreås L, Forney F, Daae FL, Torsvik V. 1997. Distribution of bacterioplankton in meromictic Lake Saelenvannet, as determined by denaturing gradient gel electrophoresis of PCR–amplified gene fragments coding for 16S rRNA. Appl Enviro Microbiol 63:3367–3373. Zhang T, Bain TS, Nevin KP, Barlett MA, Lovley DR. 2012. Anaerobic benzene oxidation by Geobacter species. Appl Environ Microbiol 78:8304–8310.

Tables

Table 1. Sampling sites and characterization of the samples

Description of Experiment Sampling sites Depth concretions

- variable sized crusts LL4A N60°01.01, E26°04.81 59 m - clay

N60°9.22, E26°36.42 55 m - spherical N60°10.55, E26°45.29 48 m - small, spherical Mixture of N60°10.63, E23°52.44 49 m - crusts concretions N60°11.48, E26°56.18 44 m - spherical N61°11.2, E18°13.81 72 m - big crusts

Table 2. Summary of PAH–degradation gene abundance, crude–oil degradation and naphthalene mineralisation in experiments with concretions and sediment. (AC= autoclaved control, NA= not applicable, ND= not determined)

14 PAH-RHD GP PAH-RHD GN Crude oil Total PAHs C naphthalene -1 -1 Oxic experiments (copies g DW) SD (copies g DW) SD disappeared (%) disappeared (%) mineralized (%) Exp. 1 Mixture of concretions 0 weeks 8.3×103 4.7×103 1.1×104 4.0×103 NA ND ND 17 weeks 3.4×104 1.6×104 2.7×104 3.5×102 59±1 ND ND AC 5.0×103 1.3×103 1.7×104 1.8×103 46 ND ND Water-oil control ND ND 0 ND ND Exp. 2 LL4A concretions 0 weeks 2.5×103 2.4×103 2.4×102 1.1×102 NA NA NA 11 weeks 2.8×103 1.9×103 2.2×106 3.4×105 70±14 97±0 7.6±0.1 11 weeks, with nutrients 4.7×102 4.4×102 2.8×105 1.8×105 81±18 92±6 5.2±0.1 Exp. 2 LL4A sediment 0 weeks 4.9×102 1.0×102 2.9×102 6.7×101 NA NA NA 11 weeks 6.6×102 4.7×102 4.8×105 3.3×105 54±16 96±1 14.2±2.8 11 weeks, with nutrients 5.4×102 3.6×102 5.2×105 1.9×105 80±8 95±1 5 .5±0.4 AC < 2.0×102 1.5×105 2.2×105 80±8 79±7 0.2 ±0.0

Anoxic experiments Exp. 1 Mixture of concretions 0 weeks 2.4×104 1.4×104 2.9×104 1.7×104 NA ND ND 17 weeks 2.3×104 2.5×104 9.1×103 1.9×104 51±6 ND ND Exp. 2 LL4A concretions 0 weeks 3.0×103 1.3×103 4.5×105 6.3×105 NA NA NA 11 weeks 1.5×104 1.7×104 2.7×106 1.0×106 61±7 97±0 1.9 ±0.3 11 weeks, with nutrients 3.3×103 2.3×103 2.4×107 2.5×107 35±31 92±6 3.5 ±0.4 Exp. 2 LL4A sediment 0 weeks 3.9×103 3.4×103 3.0×102 3.0×102 NA NA NA 11 weeks 2.9×103 1.4×103 9.2×106 1.3×107 41±26 96±1 1.3 ±0.2 11 weeks, with nutrients 5.2×103 3.3×103 9.6×105 5.9×105 47±12 95±1 2.5 ±0.4 AC < 2.0×102 8.4×101 1.4×102 10±17 32±35 0.1 ±0.0

Table 3. Closest matches (from the GenBank) of the most frequent (n>3) enriched sequences from the different treatments, and the phylogenetic classification of the sequences

Accession Similarity Phylogenetic Published number of % affiliation of the enriched enriched sequences closest match (> 97% by RDP (number of classifier)

Treatment Treatment clones) Environment closest of the match Concr HE585298.1 89-100 uncultured clone Baltic Fe-Mn Yli-Hemminki Mix (14) con59, concretions et al. 2014 Start Chromatiales, γ–Proteobacteria HE585379.1 95-100 uncultured clone Baltic Fe-Mn Yli-Hemminki (8) con185, concretions et al. 2014 Burkholderiales, β–Proteobacteria HE585324.1 87-99 uncultured clone Baltic Fe-Mn Yli-Hemminki (6) con113, concretions et al. 2014 Syntrophobactales, γ–Proteobacteria HG325295.1 94-98 uncultured, peat soil unpublished (5) Chromatiales, γ–Proteobacteria JQ177548.1 89-95 Uncultured sediment, Rotaru et al. (3) Ectothiorhodospira receiving 2012 ceae, highway run- Chromatiales, off γ–Proteobacteria JX226721.1 90-98 uncultured, deep-sea Wu et al. 2013 (3) unclassified, polymetallic δ–Proteobacteria nodules KF641574.1 89-100 uncultured, polluted soil Nielsen et al. (3) Hyphomicrobiacea 2014 e, Rhizobiales, α–Proteobacteria Concr HE585379 96-100 uncultured clone Baltic Fe-Mn Yli-Hemminki Mix (8) con185, concretions et al. 2014 week Burkholderiales, 17 oxic β–Proteobacteria HE648381 91-99 uncultured, harbour unpublished (6) β–Proteobacteria sediment HE585365 97-99 uncultured clone Baltic Fe-Mn Yli-Hemminki (4) con195, concretions et al. 2014 Rhizobiales, α–Proteobacteria

KF641612 99-100 uncultured, polluted soil Nielsen et al. (4) Rhizobiales, 2014 α–Proteobacteria HE585330 88-100 uncultured clone Baltic Fe-Mn Yli-Hemminki (3) con121, concretions et al. 2014 unclassified, δ–Proteobacteria HE585355 88-99 uncultured clone Baltic Fe-Mn Yli-Hemminki (3) con172, concretions et al. 2014 Rhodospirillaceae Rhodosirillales, α–Proteobacteria HM368347 96-99 uncultured, chemostat on Rocker et al. (3) γ–Proteobacteria humic acids 2012 Concr KC620606 96-100 uncultured, pyrite mine Hao et al. 2012 Mix (6) Albidoferax, week β–Proteobacteria 17 anoxic HE585337 90 uncultured clone Baltic Fe-Mn Yli-Hemminki (3) con136, concretions et al. 2014 Chromatiales, γ–Proteobacteria Concr HE585324 95-100 uncultured clone Baltic Fe-Mn Yli-Hemminki LL4A (11) con113, concretions et al. 2014 Start γ–Proteobacteria Oxic HE585371 99-100 uncultured clone Baltic Fe-Mn Yli-Hemminki (3) con147, concretions et al. 2014 δ–Proteobacteria JQ989595 (3) 93-99 uncultured, subseafloor unpublished α–Proteobacteria sediment AM162572 91-99 uncultured, Sepia unpublished (3) Rhodospirillaceae officinalis egg α–Proteobacteria capsule

Concr HE585320.1 96-100 uncultured clone Baltic Fe-Mn Yli-Hemminki LL4A (5) con108, concretions et al. 2014 week γ–Proteobacteria 11 oxic

Concr HE585371.1 92-100 uncultured clone Baltic Fe-Mn Yli-Hemminki LL4A (6) con147, concretions et al. 2014 Start δ–Proteobacteria Anoxic JQ989595.1 96-99 uncultured, subseafloor unpublished (3) α–Proteobacteria sediment

Concr JN977269.1 93-98 uncultured, Jiaozhao Bay unpublished LL4A (5) Rhodospirillales sediment week α–Proteobacteria 11 anoxic HE585356.1 93-100 uncultured clone Baltic Fe-Mn Yli-Hemminki (4) con173, concretions et al. 2014 δ–Proteobacteria HE585371.1 97-99 uncultured clone Baltic Fe-Mn Yli-Hemminki (3) con147, concretions et al. 2014 δ–Proteobacteria HE585379.1 91-99 uncultured clone Baltic Fe-Mn Yli-Hemminki (3) con185, concretions et al. 2014 Burkholderiales, β–Proteobacteria Sedim HE585371.1 97-99 uncultured clone Baltic Fe-Mn Yli-Hemminki Start (3) con147, concretions et al. 2014 Oxic δ–Proteobacteria Sedim FM242205.1 97 uncultured, oil- Paisse et al. week (9) γ–Proteobacteria contaminated 2010 11 oxic sediment JX016618.1 97 uncultured, marine bulk Teeling et al. (3) Cycloclasticus water 2012 γ–Proteobacteria Sedim AY216458.1 98-99 uncultured, wetland unpublished Start (3) unclassified sediment Anoxic EF459837.1 93-100 uncultured, Baltic Sea Edlund et al. (3) γ–Proteobacteria sediment 2008 Sedim FM242205.1 93-97 uncultured, oil- Paisse et al. week (5) γ–Proteobacteria contaminated 2010 11 sediment anoxic

Table S1. PAH concentrations in the oxic experiment with the concretions or sediment from the sampling station LL4A

g/kg μ g/kg 75 days μ g/kg diminished (%) days 0 μ g/kg 75 days μ g/kg diminished (%) days 0 μ g/kg 75 days μ g/kg diminished (%) days 0 μ g/kg 75 days μ Sediment0 days Concretions Sed.&nutrients Concr.&nutrients

PAH compound PAH naphtalene2-methylnapthalene1-methylnapthaleneacenapthyleneacenapthene 549.7fluorene 528.7 165.2phenanthreneanthracene 0.6fluoranthene 21.0 < 0.005 0.6pyrene 25.5benzo(a)anthracenechrysene 224.5 < 5 100 74.5benzo(b)fluoranthene 100 100benzo(k)fluoranthene < 5 5.9 30.6 5.5benzo(e)pyrene 5.4 11.9benzo(a)pyrene 677.7 2.3 270.0 3.0perylene 14.1 632.2 1.2 12.6indeno(1,2,3-cd)pyrene 2.5 11.2 5.2dibenzo(a,h)anthracene < 0.005 98 0.6 14.8benzo(g,h,i)perylene 5 < 5 < 97 0.0 5.2 10.0 20.4 < 10 59 79 < 5 54 100 56 23.9 190.8 5.7 100 5.7 5 < 7.9 4.9 66.6 100 < 63 10 < 5 28.7 181.8 3.5 < 5 4.8 4.5 10.1 445.4 5 < < 10 62 416.6 3.7 51 < 0.005 12.4 1.2 8.1 1.3 2.9 0.7 2.6 98 9.9 0.0 12.5 100 94 4.3 8.6 5 < 5 < 72 15.4 < 10 62 100 71 161.4 46 0.9 17.4 4.1 100 208.9 53.5 6.5 4.7 5 < 4.0 65 < < 5 10 23.5 5.3 460.4 3.9 9.8 < 3.5 5 < 0.005 426.5 91 67 1.6 < 10 5 < 11.6 53 11.1 0.5 1.4 3.2 2.5 97 1.6 0.0 11.3 8.2 97 5.0 8.7 53 5 5 < < < 10 67 138.1 35 14.0 55 5.2 0.4 15.5 46.2 57 5.0 6.3 5.3 22.0 5 < < 10 4.5 < 5 8.5 2.4 54 4.6 95 < 5 < 5 < 10 10.2 5 < 10.6 42 3.2 2.1 5 < 3.1 9.6 6.5 4.6 7.6 < 10 5 < 5 < 4.7 1.3 4.1 5.5 4.2 < 10 5 < < 10 5 <

Table S2. PAH concentrations in the anoxic experiment with the concretions and sediment from the sampling station LL4A

PAH compound PAH naphtalene2-methylnapthalene1-methylnapthaleneacenapthyleneacenapthene 444.4fluorene 410.8 208.0phenanthreneanthracene 0.6fluoranthene 13.7 < 0.005 < 5pyrene 14.4benzo(a)anthracenechrysene 126.6 100 < 5 43.1benzo(b)fluoranthene < 5benzo(k)fluoranthene 2.2 3.6 19.5benzo(e)pyrene 6.8 8.4 458.1benzo(a)pyrene 5.5 2.8perylene 223.6 10.2 3.0 2.4 14.1 424.0indeno(1,2,3-cd)pyrene 6.9 0.4dibenzo(a,h)anthracene 6.7 95 9.2 < 0.005benzo(g,h,i)perylene 5 5 < < 87 7.9 6.1 13.3 < 5 < 10 28 35 0 5 < 14.6 100 125.9 20 6.8 5 < 5.7 4.7 6.0 43.0 < 5 41 < 10 19.3 4.4 < 5 5.6 3.3 439.8 7.5 25 < 10 208.1 3.0 < 10 25 10.1 9.1 2.3 5 < 407.9 0.4 2.5 < 0.005 96 5.4 6.8 8.6 93 6.8 5 5.3 < 5 < 47 < 5 13.8 < 10 100 125.2 24 1.5 28 14.4 5.2 43.2 4.9 4.9 5 < 3.9 20.0 42 < 5 < 10 410.4 7.7 2.6 3.6 171.4 < 5 78 8.5 40 5.8 < 10 28 < 10 16.6 1.0 9.8 1.4 1.1 385.4 94 5 < 3.1 7.3 6.5 8.9 87 7.8 17 6.8 0.4 5 < 5 110.0 < 12.6 < 10 5 < 0 14 8.3 13.5 39.7 6.3 17.1 20.6 5.4 30 4.6 5 < 1.5 < 10 7.2 3.4 2.7 10.5 7 3.0 15.9 18 < 10 < 10 8.3 6.6 5 < 2.7 1.4 7.7 5.6 6.6 7.3 5 < 10 < 5 < 7.3 3.0 5.1 4.6 < 10 3.8 5 < < 10 < 10

Figures with captions

Fig. 1. Disappearance rates of crude oil in oxic and anoxic conditions from the experiment 1 with mixed concretions. The values are averages of triplicate flasks

Fig. 2. PAH analysis of the experiment 2 with LL4A concretion and sediment. Figures show relative proportions of 2–6 ringed PAH–compounds in the beginning and the end of the microcosm experiments. a) During the 11–week incubation in the oxic experiment the total amounts of PAH compounds decreased 97–98% in the microcosms and 80% in the autoclaved controls. b) In the anoxic experiment the total amounts of PAH compounds decreased 93–97% in the microcosms and 40% in the autoclaved controls during the 11– week incubation

Fig. 3. 14C–naphthalene mineralization rates of experiment 2 with LL4A concretions and sediment in oxic a) and anoxic b) incubations. Star indicates nutrient addition

Fig. 4. The proportions (%) of different bacterial phyla in the microcosms. Cloned sequences from the initial mixed concretions, LL4A concretions and sediments as well from the different experiments were classified by the Ribosomal Database Project Classifier

Fig. 5. The proportions (%) of different classes of Proteobacteria in the microcosms. Cloned Proteobacteria sequences from the initial concretions and sediments as well as from the different experiments were classified to different classes of Proteobacteria

Fig. 6. Environmental affiliations and the number of closest matches of the cloned sequences from the different treatments

Figure S1. Rarefaction curves showing relationship between the number of OTUs and number of sequences sampled. Rarefaction curves in the figure show individual bacterial 16S rRNA gene clone libraries from concretions collected from five different sites (mixed concretions), and two clone libraries for both concretions and sediment from site LL4A. OTU was assigned at 3% 16S rRNA sequence difference level

YEB

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