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Ecological Engineering 37 (2011) 1017– 1026

Contents lists available at ScienceDirect

Ecological Engineering

j ourna l ho me page: www.elsev ier .com/ locate /eco leng

ffect of peat re-wetting on carbon and nutrient fluxes, greenhouse gasroduction and diversity of methanogenic archaeal community

uzana Urbanová ∗, Tomás Picek, Jirí Bártaepartment of Ecosystem Biology, Faculty of Science, University of South Bohemia, Branisovská 31, 370 05 Ceské Budejovice, Czech Republic

r t i c l e i n f o

rticle history:eceived 30 October 2009eceived in revised form 11 July 2010ccepted 12 July 2010vailable online 14 August 2010

eywords:eatlande-wettingarbon dioxideethaneutrient mobilizationethanogenic archaeal community

a b s t r a c t

Many peatlands were affected by drainage in the past, and restoration of their water regime aims to bringback their original functions. The purpose of our study was to simulate re-wetting of soils of different typesof drained peatlands (bogs and minerotrophic mires, located in the Sumava Mountains, Czech Republic)under laboratory conditions (incubation for 15 weeks) and to assess possible risks of peatland waterregime restoration – especially nutrient leaching and the potentials for CO2 and CH4 production. Afterre-wetting of soils sampled from drained peatlands (simulated by anaerobic incubation) (i) phospho-rus concentration (SRP) did not change in any soil, (ii) concentration of ammonium and dissolved organicnitrogen (DON) increased, but only in a drained fen, (iii) DOC increased significantly in the drained fen anddegraded drained bog, (iv) CO2 production decreased, (v) CH4 production and the number of methanogensincreased in all soils, and (vi) archaeal methanogenic community composition was also affected by re-wetting; it differed significantly between drained and pristine fens, whereas it was more similar betweendrained and pristine bogs. Overall, the soils from fens reacted more dynamically to re-wetting than thebogs, and therefore, some nutrients (especially nitrogen) and DOC leaching may be expected from drainedfens after their water regime restoration. However, if compared to their state before restoration, ammo-

nium and phosphorus leaching should not increase and leaching of nitrates and DON should even decreaseafter restoration, especially during the vegetation season. Further, CO2 production in soils of fens and bogsshould decrease after their water regime restoration, whereas CH4 production in soils should increase.However, we cannot derive any clear conclusions about CH4 emissions from the ecosystems based on thisstudy, as they depend strongly on environmental factors and on the actual activity of methanotrophs insitu.

© 2010 Elsevier B.V. All rights reserved.

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. Introduction

Re-wetting of peatlands is currently a wide spread methodor restoration of wetland ecosystems which were drained due tontensification of forestry and agriculture. Almost 70% of peatlandsn the Sumava Mountains (Bohemian Forest, Czech Republic) werenfluenced by drainage in the past and they are now being restored

ithin the long-term project “Programme of Peatland Restorationn the Sumava National Park”.

Under natural conditions, peatlands play an important rolen the global C cycle as a long-term sink of atmospheric C and

source of CH4 (Gorham, 1991) and they have a high poten-ial for nutrient retention and nutrient transport (Mitsch andosselink, 2000). Peatland restoration aims to bring back these

∗ Corresponding author. Tel.: +420 387 772 261; fax: +420 387 772 368.E-mail address: urbanz00@prf.jcu.cz (Z. Urbanová).

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925-8574/$ – see front matter © 2010 Elsevier B.V. All rights reserved.oi:10.1016/j.ecoleng.2010.07.012

atural functions (Vasander et al., 2003); however, the successf restoration is affected by many factors, especially by the statef degradation and type of peatland (Schumann and Joosten,006).

Drainage also brings about changes in the chemistry of the sur-ace peat, which can have an impact on nutrient availability tolants (Aerts et al., 2006) and affect the quality of surface and runoffater (Lundin and Bergquist, 1990; Prevost et al., 1999). Aeration of

he peat and subsequent mineralization and nitrification of organic result in an increase of NO3

− concentrations in the pore water ofrained peatlands (Olde Venterink et al., 2002; Holden et al., 2004).hese processes are intensified by higher groundwater level fluc-uation, which is typical for degraded peatlands due to the alteredydraulic properties of the peat (Price et al., 2003; Tiemeyer et al.,

006). The aim of re-wetting is to reduce soil aeration and decrease

mineralization, and hence to decrease N availability for wetlandlants. Olde Venterink et al. (2002) showed that re-wetting of driedoil cores can strongly stimulate denitrification but N mineraliza-

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ion did not decrease significantly. Phosphorus (P) is frequentlyorbed to Fe- or Al-hydroxides in aerobic condition and becomesemporarily immobilized (Zak et al., 2004). Re-wetting can leado enhancement of P mobilization due to the reduction of theseomplexes under anaerobic soil conditions and low redox poten-ial (Olde Venterink et al., 2002; Tiemeyer, 2007). In general, theisk of water pollution by elevated solute concentrations dependsainly on the hydrological conditions of the area after drainage or

e-wetting (Price et al., 2003).The production of CO2 and CH4 depends on the position of

he water level, soil temperature, microbial activity in the peat,lant community structure and the chemical characteristics of peatBubier et al., 1993; Whiting and Chanton, 1993; Yavitt et al., 1997).rainage causes simultaneous changes in vegetation and decom-osition processes, and hence greenhouse gas fluxes in the mire.rained peatlands tend to emit more CO2 (Moore and Dalva, 1993;ilvola et al., 1996), while emissions of CH4 greatly decrease dueo limited anaerobic conditions (Nykänen et al., 1995; Minkkinent al., 2002). The actual results of restorations are encouraging,ecause they reflect the positive effect of restoration of both fennd bog sites in initiating vegetation succession and carbon bal-nce development towards those of pristine mires. Most of studiesere conducted on cut-away peatlands, where decreasing total res-iration and higher incorporation of CO2 to the system due to the

ncrease of plant cover was observed after re-wetting (Tuittila et al.,999). The C balance after re-wetting is strongly influenced by veg-tation structure (Komulainen et al., 1999; Kivimäki et al., 2008).omulainen et al. (1998) and Tuittila et al. (2000) observed that aigher water level after re-wetting was followed by increased CH4missions. Nevertheless, CH4 emissions from restored peatlandsay remain at a lower level for a long period of time until they

ave become fully vegetated by mire plants.A significant relationship has been described between

ethanogen community composition and environmental char-cteristics such as vegetation type and temperature, whichnfluence both methanogen community activity and dynamicsRooney-Varga et al., 2007). Several studies confirmed variation inhe methanogen community between peatlands types (bog, fen)elated to vegetation (Basiliko et al., 2003; Galand et al., 2005;avitt et al., 2005). Methanogen communities have also beenescribed in undisturbed peatlands, as well as the effect of short-nd long-term drought on microbial diversity (Fenner et al., 2005;aatinen et al., 2007; Kim et al., 2008), but the direct effect ofrainage and restoration on methane-producing archaea is still aatter of debate.Although there have been several studies about the effect

f re-wetting on nutrient mobilization (leaching) and microbialrocesses, including gas emissions or production, their conclu-ions are not unambiguous. Understanding ecosystem functioningnd reaction to re-wetting require consideration of both theydrology and soil processes together with vegetation struc-ure, as well as their interactions. A comprehensive study ofutrient losses, and CO2 and CH4, production and changes ofiversity of archaeal methanogenic community, will help to antic-

pate how peatlands will function after restoration of their wateregime.

The purpose of our study was to simulate re-wetting of soilsf different types of drained peatlands (bogs and minerotrophicires) in a laboratory incubation experiment and assess possible

isks of peatland water regime restoration (especially its effect onutrient leaching and CH4 production potential). To accomplish

his, we measured CO2 and CH4 rates production, nutrient releasend immobilization, the number of methanogens and the diversityf the archaeal methanogenic community in peat samples underaboratory conditions.

WcEb

eering 37 (2011) 1017– 1026

. Materials and methods

.1. Study sites

Three ombrotrophic bogs and two minerotrophic mires in dif-erent states of degradation were chosen for peat sampling asepresentative of two main types of peatland in the Sumava Moun-ains (Bohemian Forest):

. intact bog (BOG),

. drained bog (BOGdr1),

. degraded drained bog (BOGdr2),

. intact minerotrophic mire (FEN), and

. drained minerotrophic mire (FENdr).

These sites are located in the Sumava National Park in the south-rn part of the Czech Republic. All of these study sites are includedn the long-term project “Programme of peatland restoration in theumava National Park”. Two ombrotrophic bogs, Blatenska (BOG)nd Schachtenfilz (BOGdr1), are located on the central plateau ofhe Sumava Mountains, which contains frequent patterned miresnd extensive waterlogged spruce forests at an average altitude ofbout 1150 m a.s.l. The cold and humid climate is characterized by

total annual precipitation up to 1400 mm and mean annual tem-erature of 3.2 ◦C. The other three study sites are situated in theollow of the Kremelna River at an average altitude of about 850 m.s.l. The mean annual temperature is less than 4 ◦C and mean totalrecipitation up to 1100 mm.

Vegetation structure of each site reflects the hydrological con-itions. The surface of the intact bog (BOG) is structured withenerous hollows (Leuco-Scheuchzerion association), lawns withhe dominant Trichophorum caespitosum and hummocks withndromeda polifolia. Shrubs occur only on the margin, which is sur-ounded by Pinus × pseudopumilio and waterlogged spruce forest.n the medium disturbed site (BOGdr1), the lawns with dominant. caespitosum are preserved only in the more hydrologically stableentral part (between ditches) and hollows are missing. Ditchesre surrounded by shrubs (Vaccinium myrtillus, Vaccinium uligi-osum) and expansion of Picea abies is clearly evident on the wholeite. The degraded drained site (BOGdr2) has much more changedegetation structure with dominant shrubs V. uliginosum in theentral part and V. myrtillus along the margin. Molinia caeruleaxpanded on to the whole site together with Betula pubescens.he advanced successional stage of degradation on BOGdr2 isrobably caused by earlier drainage compared to BOGdr1. The

ntact minerotrophic mire (FEN) is covered by typical vegetationith dominant Carex rostrata and almost 100% Sphagnum cover.

he drained minerotrophic mire (FENdr) is composed of differentegradation states depending on the intensity of drainage, fromell preserved parts with C. rostrata to a dry part with dominant

arex nigra, Carex brizoides, M. caerulea, Nardus stricta and otherraminoids.

These peatlands have been drained namely for forestry pur-oses since the 19th century with local intensive drainage madeuring the last 50 years. In the case of the drained fen, the pur-ose of drainage was most probably to increase hay production.o fertilization was done after drainage at all sites.

.2. Environmental parameters

Characteristics of the five study sites are shown in Table 1.

ater table was measured manually (at two week interval) or

ontinuously using dataloggers with a water level sensor (Fiedler,lectronics for Ecology, Czech Republic) in minimally six plasticoreholes at each site. From two to four samples of pore water

Z. Urbanová et al. / Ecological Engineering 37 (2011) 1017– 1026 1019

Table 1Physico-chemical characteristics of peat and pore water sampled at five study sites (average ± standard deviations). Peat characteristics were measured in the upper layer(0–300 mm), pore water was sampled from perforated pipes in the layer 0–1000 mm.

Study site FEN FENdr BOG BOGdr1 BOGdr2State Intact minerotrophic

mireDrained minerotrophicmire

Intact bog Drained bog Degraded drainedbog

Peat characteristicsBulk density [g cm−3] 0.16 ± 0.02 0.92 ± 0.30 0.30 ± 0.06 0.26 ± 0.07 0.36 ± 0.07Total C [%] 43.4 ± 0.2 32.8 ± 4.3 47.9 ± 1.5 48.2 ± 0.6 48.4 ± 1.4Total N [%] 1.72 ± 0.12 1.50 ± 0.18 1.86 ± 0.39 1.98 ± 0.11 1.42 ± 0.10C–N ratio 25.4 ± 1.8 22.0 ± 1.6 26.9 ± 5.2 24.4 ± 1.5 34.4 ± 2.6Total P [%] 0.10 ± 0.01 0.13 ± 0.01 0.06 ± 0.01 0.08 ± 0.01 0.06 ± 0.00Total S [%] 0.19 ± 0.06 0.11 ± 0.02 0.06 ± 0.02 0.18 ± 0.04 0.13 ± 0.01

Pore water characteristicsWater table [cm] −8.25 ± 0.66 −33.4 ± 10.4 −12.5 ± 9.0 −24.0 ± 13.1 −25.4 ± 14.0pH 5.5 ± 0.07 4.6 ± 0.1 4.0 ± 0.1 3.9 ± 0.2 4.1 ± 0.1Conductivity [�S m−2] 42.3 ± 4.4 31.2 ± 11.9 13.2 ± 7.7 14.2 ± 5.1 26.5 ± 5.1NH4

+ [mg L−1] 0.33 ± 0.22 0.70 ± 0.46 0.13 ± 0.10 0.34 ± 0.27 0.23 ± 0.12NO3

− [mg L−1] 0.04 ± 0.14 0.04 ± 0.12 0.01 ± 0.06 0.04 ± 0.10 0.02 ± 0.04SO4

2− [mg L−1] 1.62 ± 0.73 0.84 ± 0.69 0.22 ± 0.17 0.34 ± 0.24 0.41 ± 0.35SRP [mg L−1] 0.09 ± 0.08 0.57 ± 0.58 0.03 ± 0.03 0.04 ± 0.03 0.19 ± 0.14Al [�g L−1] 2.05 ± 2.43 2.16 ± 3.03 0.18 ± 0.12 0.15 ± 0.08 0.47 ± 0.21Ca2

+ [�g L−1] 1.77 ± 1.69 1.45 ± 0.71 0.29 ± 0.40 0.47 ± 0.42 1.24 ± 1.02+ −1

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Fe2 [�g L ] 3.43 ± 2.04 2.65 ± 3.41

Mg2+ [�g L−1] 1.21 ± 0.81 2.13 ± 2.43

DOC [�g L−1] 17.24 ± 9.05 37.50 ± 6.02

ere taken monthly during the vegetation season (V-X) for detailedydro chemical analysis of each site, including content of mainations and anions (SO4, NO3, NH4, PO4, Ca, Mg, Al, Fe), pH, con-uctivity and DOC.

.3. Soil sampling and analysis

Soils were sampled in October 2008. At each site, eight samplesere collected using a corer from the upper layer of soil (300 mm)

each sample consisted of 10 subsamples mixed together). Soilsere then sieved through a 5-mm mesh and stored at 4 ◦C until

nalyzes (laboratory experiment).The same soil samples were analyzed for total C, N and S con-

ent using a Micro-cube elemental analyzer (Elementar, Germany).otal phosphorus was determined by the perchlorate mineraliza-ion method.

.4. Laboratory experiment

Homogenized (sieved) soil samples were incubated under labo-atory conditions at 15 ◦C in two replicates for 100 days. Ten gramsf soil were weighed into glass bottles (100 ml) and stopperedirtight by rubber stoppers. Half of the samples were incubatednder aerobic conditions (air in headspace) and half of the samplesnder anaerobic ones (10 ml of distilled water were added to eachnaerobic flask and then the headspace was flushed with helium).he aerobic bottles were regularly ventilated to keep conditionsnside the flask aerobic. The anaerobic bottles were closed tightlyor the whole incubation period and were submerged into boxesith water to prevent oxygen diffusion into the samples. Gases in

he headspace were sampled using a syringe (200 �l) through aubber stopper.

Concentration of CO2 and CH4 in the headspace of the incuba-ion flasks were measured regularly (at one or two week intervals)uring whole experiment. CO2 was determined using an HP 6850as chromatograph (Agilent, USA) equipped with a 0.53 mm × 15 m

P-Plot Q column and a 0.53 mm × 15 m HP-Plot Molecular SieveA column, and a thermal conductivity detector, using heliums the carrier gas. CH4 was determined using an HP 6890 gashromatograph (Agilent, USA) equipped with a 0.53 mm × 30 m GS-

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0.36 ± 0.24 0.39 ± 0.16 0.39 ± 0.150.10 ± 0.21 0.09 ± 0.13 0.65 ± 0.84

35.55 ± 20.28 35.42 ± 7.76 53.92 ± 10.90

lumina column and a flame ionization detector, using nitrogen ashe carrier gas.

At the start (day 0) and end of the experiment (day 100), dis-olved organic carbon (DOC), total soluble nitrogen (TSN), solubleeactive phosphorus (SRP), nitrates and ammonium were mea-ured in soil solution after centrifugation (5000 × g) and filtrationhrough glass fibre filters (0.45 mm, MN-GF5, Germany). pH waslso measured in the same samples but before their centrifuga-ion and filtration. Concentrations of DOC and TSN in soil solutionere analyzed on a LiquiTOC II (Elementar, Germany). Concentra-

ions of NH4-N, NO3-N and SRP were measured on flow injectionnalyzer (FIA Lachat QC8500, Lachat Instruments, USA). DON (dis-olved organic nitrogen) was calculated by subtracting the sum ofhe mineral forms of nitrogen from TSN (total soluble nitrogen).

.5. Extraction of DNA

A duplicate of each soil sample (0.25 g) was taken for DNAxtraction at the start and at the end of anaerobic incubation. Theower Soil DNA Isolation kit (MoBio Laboratories Inc., Carlsbad, CA,SA) was used for isolation of genomic DNA from soil according

o manufactures instructions with some modifications. Mini Bead-eater (BioSpec Products, Inc.) at the speed of 6 m s−1 for 45 s wassed for better disruption of cell walls. DNA was stored in 1.5 mlppendorf microtubes in a freezer (−20 ◦C) until analyses.

.5.1. Composition of methanogenic Archae (PCR-DGGE analyses)The communities of methanogenic Archae were analyzed using

G357F (5′-CCC TAC GGG GCG CAG CAG-3′) and MG0691R (5′-GGATA CAR GAT TTC AC-3′) primer pairs to amplify the 367 bp frag-ent of the archaeal 16S rDNA by PCR (Watanabe et al., 2004).

40 bp GC-clamp (5′-CGC CCG CCG CGC GCG GCG GGC GGGCG GGG GCA CGG GGG G-3′) was attached to the 5′ end of theG357F primer to avoid complete separation of DNA strands dur-

ng denaturing electrophoresis. The reaction medium consisted of �l of PCR buffer (Roche, 100 mM Tris–HCl, pH 8.3, 500 mM KCl,5 mM MgCl2), 1 �l of dNTP (10 mM), 1 �M of each primer, 1 �l

f GC-rich solution for enhancing PCR efficiency (Roche, France),00 ng of Bovine Serum Albumin (Fermentas, Italy), 0.5 �l of Taq-olymerase (5 Units/�l, Roche, France) and 2 �l of genomic DNA10 ng) brought to a final volume of 50 �l. The amplification pro-

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ramme consisted of an initial cycle of denaturation at 94 ◦C for min followed by 35 cycles of denaturation at 94 ◦C for 60 s,nnealing at 53 ◦C for 60 s and extension at 72 ◦C for 120 s. Themplification concluded with a final elongation step at 72 ◦C for

min. Size of PCR amplicons was verified using 1.5% (w/v) agarosend electrophoresed with specific markers.

For DGGE analyses, 10 �l of PCR products (approximately50 ng) were loaded into a polyacrylamide gel (8%) with a 30–60%enaturing gradient for 16 h at 100 V and 60 ◦C using the DCodeTMniversal Mutation Detection System (BioRad, USA). The 100%enaturing stock solution consisted of 105 g urea and 100 mleionised formamide/250 ml. Gels were stained with SYBR®Green1:10000) and visualized under UV light.

Bands in acrylamide gels were identified and analyzed using theiorad Quantity One software v. 4.5.2. The band position toleranceas set at 1% and background subtraction was applied. Within-gel

omparisons, taking into account the presence/absence of bands,ere performed using multivariate analyses.

.5.2. Quantification of methanogenic Archae (qPCR assay)The gene (mrcA) coding coenzyme M was used for quantitative

nalysis of methanogens. Quantitative PCR was performed withhe primer pair ME1 (5′-GCM ATG CAR ATH GGW ATG TC-3′) and

CR1R (5′-ARC CAD ATY TGR TCR TA-3′) as described by Halest al. (1996). Amplification was carried out with an ABI Step OneApplied Biosystems) by using SYBR green as the detection systemn a reaction mixture of 20 �l containing 0.5 �M (each) primer;0 �l of Power SYBR®Green PCR master mix, including AmpliTaqold® DNA Polymerase, Power SYBR®Green PCR buffer, deoxynu-leoside triphosphate mix with dUTP, SYBR green I, ROX, and 5 mMgCl2 (Power SYBR®Green PCR Master Mix; Applied Biosystems,SA); and 2 �l of template DNA corresponding to 10 ng of totalNA. Bovine serum albumin (BSA, 500 ng/reaction; Fermentas,

taly) and dimethylsulfoxide (DMSO, 12.5 �mol/reaction) was usedo enhance PCR efficiency. Thermal cycling conditions for the mrcAene was performed as described by Kim et al. (2008) with someodifications: 40 cycles with denaturation at 95 ◦C for 30 s, primer

nnealing at 60 ◦C for 60 s, and extension at 72 ◦C for 60 s. Fluores-ence was measured after each extension step. Melting curve andgarose electrophoresis (1.5% w/v, 110 V, 45 min) was performedor quality verification of the PCR product. Thermal cycling, fluo-escent data collection, and data analysis were carried out with theBI Step One. Two independent quantitative PCRs were performed

or each soil replicate. Standard curves were obtained with seriallasmid dilutions of a known amount of plasmid DNA containing aragment of the mrcA gene.

.6. Statistical evaluation and multivariate analyses

Nested design ANOVAs were used to test for differences amongreatment types at each locality, where treatment was nested inocality. The statistical significant differences between treatmentseffect of aerobic and anaerobic conditions) were examined forxtractable nutrients and pH using one-way ANOVA (p < 0.05).ates of gases production were evaluated using repeated measuresNOVA. All statistical analyses were conducted using STATISTICA

(StatSoft Inc., USA).A unimodal type of constrained ordination, canonical analy-

is (CCA), was used to evaluate the relation between the soilhemical properties (explanatory variables) vs. composition of therchaeal methanogenic communities (response variables) at the

nd of incubation. The contribution of each explanatory variableas tested using forward selection and the Monte-Carlo per-utation test within the CCA framework (p < 0.05). Only those

xplanatory variables that showed significant marginal and con-

a7

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eering 37 (2011) 1017– 1026

itional effects on the microbial communities were includedn the diagrams. The results were summarized using biplotiagrams. The distance between archaeal methanogenic com-unities illustrates the magnitude of differences between the

ites. The relative length and position of arrows show the extentnd direction of response of archaeal methanogenic communitytructures to the selected soil chemical properties. The analysisas performed using CANOCO ver. 4.5 (ter Braak and Smilauer,

998).

. Results

.1. Effect of re-wetting on nutrients mobilization/immobilization

Concentrations of NH4+ and NO3

− increased under aerobic con-itions in all samples (Fig. 1a and c). The concentration of NO3

−

as significantly lower under anaerobic conditions compared toerobic ones (F(1.14) = 26.2; p < 0.01), but the significant decreasen NO3

+ concentrations compared to beginning values was onlyor FENdr and BOGdr2. However, NH4

+ concentrations slightlyncreased in all samples except BOGdr2 under anaerobic conditionsF(1.14) = 4.6; p < 0.05) (Fig. 1b and d). Concentration of dissolvedrganic nitrogen (DON) either decreased or remained the samender both aerobic and anaerobic conditions in all samples exceptENdr, where concentration of DON increased significantly underoth aerobic and anaerobic conditions.

Effect of aerobic or anaerobic incubation on soluble reactivehosphorus (SRP) concentration was almost negligible (not signif-

cant, p > 0.05). The SRP concentration slightly decreased for FENnd BOG, while it increased slightly for BOGdr1, BOGdr2 and FENdrnder both aerobic and anaerobic conditions. The concentration ofRP measured in our experiment correlated neither with the con-entration of total P in peat or with SRP measured in pore water initu.

Under aerobic conditions, concentration of dissolved organicarbon (DOC) decreased significantly in all samples except BOGwithout change) (Table 2) and increased (for FEN, FENdr, BOGdr2)r remained the same (for BOG, BOGdr1) under anaerobic condi-ions.

pH decreased after aerobic incubation in all samples except BOG,here the changes were negligible (Table 2). Under anaerobic con-itions, pH increased for FEN and FENdr but remained the same forOG, BOGdr1 and BOGdr2.

Overall, the observed changes of all chemical parameters tendedo be much more pronounced for samples from minerotrophic

ires (FEN, FENdr) compared to samples from the bog sites.

.2. Effect of re-wetting on CO2 and CH4 production

Aerobic CO2 production rates ranged from 0.08 to.96 �mol CO2 g−1 h−1 (Table 3). The highest CO2 productionate was measured for FEN and the lowest for FENdr. The rate ofO2 production was significantly lower under anaerobic conditionss compared to aerobic ones (F = 3.5680; p < 0.01). The rate of CO2roduction decreased at the start of the experiment (days 5–20)nder both aerobic and anaerobic conditions, respectively (Fig. 2).hen it was more or less constant or slightly increasing undererobic conditions while it decreased during anaerobic incubation.naerobic CO2 production rates followed a similar pattern to theerobic rates; there was a strong correlation between aerobic andnaerobic CO2 production rates (r = 0.88, Fig. 3a). Ratios between

erobic and anaerobic CO2 production rates ranged from 2.0:1 to.4:1, with a mean of 4.7:1 (Table 3).

Average CH4 production rate (anaerobic incubation) rangedrom 0.0003 to 0.22 �mol CH4 g−1 h−1 (Table 3). The rate of CH4

Z. Urbanová et al. / Ecological Engineering 37 (2011) 1017– 1026 1021

Fig. 1. Changes in concentration of nutrients (NH4, NO3, DON and SRP, �g g−1) in peat from pristine fen and bog (a) after aerobic laboratory incubation, (b) after anaerobiclaboratory incubation, in peat from drained fen and drained bogs, (c) after aerobic laboratory incubation and (d) after anaerobic laboratory incubation (average ± standarddeviations, n = 8).

Table 2pH and concentration of dissolved organic carbon (DOC) at the start of laboratory experiment and after aerobic and anaerobic incubation (average ± standard deviations,n = 8). Different letters in columns indicate statistically significant differences between treatments (p < 0.05).

Study site FEN FENdr BOG BOGdr1 BOGdr2

pH Begining 5.00 ± 0.32a 4.11 ± 0.54a 4.07 ± 0.28a 3.78 ± 0.27a 3.90 ± 0.20a

Aerobic 4.75 ± 0.39ab 3.69 ± 0.63b 4.15 ± 0.35a 3.60 ± 0.15ab 3.79 ± 0.22b

Anaerobic 5.32 ± 0.30ac 4.93 ± 0.33c 4.08 ± 0.16a 3.92 ± 0.09ac 3.96 ± 0.14ac

DOC [�g g−1] Begining 838 ± 208a 204 ± 65a 1465 ± 640a 1065 ± 485a 733 ± 192a

± 52 ± 22

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Aerobic 395 ± 259b 160Anaerobic 2240 ± 294c 1183

roduction increased in all samples from the start of the exper-ment (Fig. 4). The rates of CH4 production were significantly

ower for the drained sites BOGdr1, BOGdr2 and FENdr as com-ared to the intact sites. The highest rate of CH4 productionas measured for FEN (0.22 ± 0.06 �mol CH4 g−1 h−1), where 50%

f the C mineralized was found in the form of CH4 and 50%

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able 3verage rate of CO2 production under aerobic conditions (aerobic CO2) and under anaero

o anaerobic CO2 production rate (R CO2), average rate of CH4 production (CH4, �mol gethane production rate (R CO2–CH4) during laboratory experiment (average ± standard

Study site FEN FENdr

Aerobic CO2 0.96 ± 0.36 0.08 ± 0.03

Anaerobic CO2 0.23 ± 0.14 0.04 ± 0.02

R CO2 4.2 2.0

CH4 0.22 ± 0.14 0.0003 ± 0.0009

R CO2–CH4 1.0 133.3

b 1606 ± 1317a 762 ± 217b 255 ± 54b

2c 1630 ± 562a 1176 ± 227ac 980 ± 417c

n the form of CO2 after anaerobic incubation. Ratios betweennaerobic CO2 and CH4 production rates ranged from 1.1 to

39.8, with a mean of 59.0 (Table 3). There was a strong corre-

ation between anaerobic CH4 production rate and aerobic andnaerobic CO2 production rates (r = 0.84 and 0.98, respectively,ig. 3b).

bic conditions (anaerobic CO2; �mol g−1 h−1), ratio of aerobic CO2 production rate−1 h−1) under anaerobic conditions and ratio of anaerobic CO2 production rate to

deviations, n = 8).

BOG BOGdr1 BOGdr2

0.39 ± 0.22 0.28 ± 0.10 0.15 ± 0.040.05 ± 0.06 0.05 ± 0.04 0.04 ± 0.037.8 5.6 3.80.0060 ± 0.0184 0.0006 ± 0.0008 0.0006 ± 0.00138.3 83.3 66.7

1022 Z. Urbanová et al. / Ecological Engineering 37 (2011) 1017– 1026

Fig. 2. The rate of CO2 production (�mol CO2 g−1 h−1) (a) under aerobic con-d(

usoaT

Fd

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3c

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tabimsity index (H = 0.7 ± 0.0).

Fa

itions and (b) under anaerobic conditions during the laboratory incubationaverage ± standard deviations, n = 8).

The rates of CO2 and CH4 production were also recalculated perpper 20 cm of peat per square meter using bulk densities for eachample to show what could be the situation in situ. The highest aer-

bic CO2 production rate was for FEN then for BOG, FENdr, BOGdr1nd BOGdr2 (values ranged from 29.9 to 10.6 mmol CO2 m2 h−1).he highest potential values for anaerobic CO2 respiration rate

as

ig. 3. The relationship between aerobic CO2 production rate and anaerobic CO2 productind aerobic CO2 production rate.

ig. 4. The rate of CH4 production (�mol CH4 g−1 h−1) under anaerobic conditionsuring laboratory incubation (average ± standard deviations, n = 8).

ere calculated for FENdr then for FEN, BOG, BOGdr2 and BOGdr17.5, 7.1, 3.1, 2.7 and 2.4 mmol CO2 m2 h−1, respectively). PotentialH4 production rate was also the highest for FEN and BOG anduch lower for FENdr, BOGdr1 and BOGdr2.

.3. Effect of re-wetting on microbial methanogenic archaealommunity

The number of methanogens was significantly influenced bye-wetting (F(1.18) = 5.0269; p < 0.05) despite of high variabilityetween sites. The number of methanogens increased in all samplesuring anaerobic incubation with the highest increase measuredor FENdr (Fig. 5). However, no correlation was found betweenumber of methanogens and the rate of CH4 production.

The DGGE profiles showed quite diverse fingerprints ofhe methanogenic archaeal communities in the five peat soilsfter anaerobic incubation (Fig. 6). FEN had the highest num-er of species/bands (11) and the highest Shannon diversity

ndex (H = 2.2 ± 0.2) while BOGdr1 had the lowest number ofethanogenic archaeal species (2) and the lowest Shannon diver-

The first and second canonical axes explained 47.2% of the vari-bility of the methanogen communities in the five studied soilamples (Fig. 7). The FEN and FENdr methanogen communities

on rate (a). The relationship between anaerobic CH4 production rate and anaerobic

Z. Urbanová et al. / Ecological Engineering 37 (2011) 1017– 1026 1023

Fmb

wBesnTtFl(

4

4

i(2nl

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Fig. 7. Ordination biplot of CCA displaying the effect of environmental variables(full line arrows) explaining the variance in methanogenic archaeal communitycomposition after the incubation experiment (two samples of peat from each studys(c

oSicWpadlNoma

ig. 5. Effect of re-wetting on number of methanogens. The bars show number ofethanogens (gen copies g−1 soil) before (START) and after (END) anaerobic incu-

ation (average ± standard deviations, n = 2).

ere quite diverse from the rest, while the BOG, BOGdr1, andOGdr2 methanogen communities showed higher similarity. Sixnvironmental variables (pH, NH4

+, SRP, DOC, TSN and NO3−) were

elected for multivariate analyses. Only pH and DOC showed sig-ificant marginal effect (p < 0.05) on the methanogen communities.he variance explained by pH and DOC was 32.2% and 29.4%, respec-ively. pH and DOC had a strong effect on methanogens from theEN site, while the effect on methanogens from the other sites wasower and decreased in the order BOG, BOGdr2, BOGdr1 and FENdrFig. 7).

. Discussion

.1. Nutrients mobilization/immobilization

Drainage has been shown to significantly influence N mineral-zation and its subsequent availability for plants in many studies

Prevost et al., 1999; Olde Venterink et al., 2002; Tiemeyer et al.,007); however, the effect of re-wetting on N mineralization hasot been unambiguously defined. Re-wetting can strongly stimu-

ate denitrification, which can play an important role in the removal

ig. 6. DGGE profile of methanogenic Archaea from five different sites measured inuplicates.

tal

iaNad

tttGpft((

csBww

ite). Monte-Carlo permutation test was calculated and two environmental variablespH and DOC) explaining the most variability of methanogenic archaeal communityomposition are shown.

f N from restored peatlands (Silvan et al., 2002; Kieckbusch andchrautzer, 2007). We did not measure denitrification rate directlyn our experiment but we observed a significant decrease in NO3

−

oncentration for the FENdr sample during anaerobic incubation.e assume that denitrification was probably the most responsible

rocess for NO3− removal. After re-wetting of drained peatlands

much lower rate of nitrification can be expected and thereforeenitrification will become a less important process at restored

ocalities. Most mineral N will be present either in the form ofH4

+ or organic forms. Olde Venterink et al. (2002) also pointedut that denitrification rates under laboratory conditions can beuch higher than rates measured in the field. In our case, we can

lso expect a lower rate of denitrification after re-wetting due tohe vegetation present on our study sites. In field conditions, NO3

−

nd NH4+ are used by plants and therefore denitrification can be

imited by low NO3− concentration.

The effect of re-wetting on organic N compounds can be eithernsignificant (Olde Venterink et al., 2002), or peatlands can act as

source of DON in the first years after re-wetting and the total balance can be negative (Kieckbusch and Schrautzer, 2007). Welso measured significantly enhanced concentration of DON in therained minerotrophic mire (FENdr) after re-wetting.

The potential for N leaching from restored peatlands in cen-ral Europe is higher as compared to northern peatlands as theotal N contents in Central European peats (soil upper layer from 0o 350 mm) are usually higher than in Fennoscandia; e.g. Zak andelbrecht (2007) measured from 2.7 to 3.4% of total N in peats sam-led from a fen in northeastern Germany, Tiemeyer et al. (2007)ound 2.23–2.80% total N in a drained peatland in Germany andotal N content ranged from 1.42 to 1.98% for our five study sitesTable 1). For ombrotrophic bogs in Finland, e.g. Peltoniemi et al.2009) documented values ranging from 0.84 to 1.48%.

Re-wetting did not affect P availability in our experiment. Thehanges in SRP concentration were negligible and we observed a

light increase of SRP concentration for all drained sites (FENdr,OGdr1, BOGdr2) after anaerobic incubation. Therefore, in our case,e can suppose a low risk of enhanced P availability after in situ re-etting. On the contrary, other studies (Olde Venterink et al., 2002;

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024 Z. Urbanová et al. / Ecological

ak et al., 2004; Tiemeyer et al., 2005; Kieckbusch and Schrautzer,007) showed an increased extractable pool of P in re-wetted soils.xport of P to the surface water can be influenced by specific hydro-hemical conditions of the site. These studies also supposed thatydrochemical conditions become more stable during the years fol-

owing re-wetting and the mobile nutrient pool probably decreasesue to the intensification of processes such as sedimentation andeat formation. Most of the studies mentioned above were done oneatlands with high trophic status affected by agricultural activi-ies (rich fens, wet meadows). Therefore, we can suppose that theotal P pool was higher on these sites as compared to our studyites (Kellogg and Bridgham, 2003), which had never been used forgriculture (drained bogs) or had been used only as an extensiveasture (drained minerotrophic mire). The mechanisms control-

ing cycling and mobility of P also differ between different types ofeatlands; this is especially when comparing bogs and mezotrophicens. There the availability of P is controlled not only by hydro-hemical conditions but also by vegetation and microorganismsKellogg and Bridgham, 2003). We can also assume that microor-anisms and vegetation will quickly assimilate available P releasedfter re-wetting. On the other hand, the effect of temperature maye important under in situ conditions and may cause some unex-ected nutrient leaching. For example, Koerselman et al. (1993)ound that freezing and thawing of peat caused some nutrienteaching from their sites. Our experiment was done under constantemperature (15 ◦C), so we cannot say what will happen in the fieldfter thawing of the upper peat layer in spring. From our experi-ent, we can at least predict that some nutrients will be released

rom peat in early spring.In our experiment, re-wetting was followed by increased DOC

oncentration in some samples (FEN, FENdr, BOGdr2) or DOC con-entration remained the same (BOG, BOGdr1). Although the effectf restoration on DOC dynamics has not been clearly defined yet,allage et al. (2006) described drain blocking of a blanket peatland

s a very successful technique in reducing DOC concentration andoss. We can suppose that restoration could lead to increased DOConcentration during the first years but the mobile pool of DOCould decrease with more stable conditions during the followingears after restoration.

Among other chemical factors, pH is one of the most impor-ant affected by re-wetting. In our experiment, pH did not changeignificantly for bogs, but it increased by one unit in the draineden after 15 weeks of anaerobic incubation. It reached the value of.9 which is almost the same as for the pristine fen at the begin-ing of experiment (5.0). The increase of pH may increase rates of

mportant microbial processes, e.g. methanogenesis (Wang et al.,993) or nitrification (Ste-Marie and Pare, 1999) and so affect DOCr nutrient leaching.

.2. CO2 and CH4 production

Aerobic CO2 production rates ranged from 0.08 to.96 �mol CO2 g−1 h−1, whereas anaerobic CO2 productionates were significantly lower and ranged from 0.04 to.23 �mol CO2 g−1 h−1. Similar CO2 production rates were alsoeasured by other authors (Moore and Dalva, 1993, 1997; Bergman

t al., 1999; Glatzel et al., 2004). We observed significantly lowererobic and anaerobic CO2 production rates for the samples fromhe drained fen compared to the intact fen, while the samples fromhe drained and the intact bogs showed lower variation in CO2roduction rates between sites. These results confirm the concept

f more dynamic changes and degradation of minerotrophic miress compared to bogs after drainage (Minkkinen et al., 1999).

The average ratio of aerobic to anaerobic CO2 production rateeasured in our study (4.7:1) was similar to those reported by oth-

bonm

eering 37 (2011) 1017– 1026

rs for short-term laboratory incubations of peat (Moore and Dalva,993, 1997; Bergman et al., 1999; Glatzel et al., 2004). Limitation ofnaerobic CO2 production rate by accumulation of metabolites andcidic conditions as mentioned by Magnusson (1993) and Bergmant al. (1999) for long-term incubations was not proven in our exper-ment.

The strong correlation between aerobic and anaerobic CO2 andnaerobic CH4 production rates suggests, that in bog samples,ethanogenic bacteria could be limited by low concentration of

vailable (easily decomposable) organic material. In our study,reater availability of substrate for soil microorganisms was indi-ated by the large potential CO2 and CH4 production rates foramples from the intact minerotrophic mire (FEN) as comparedo samples from bogs and drained minerotrophic mire. Similaresults were described by Moore and Dalva (1997), who observedhe highest CO2 production rate for peat with herbaceous originrelatively easily decomposable) than either bryophyte or ligneousrigin (more resistant to decomposition). The decomposabilityf organic matter (peat) may be explained by the degree of itsumification (Nilsson and Bohlin, 1993; Glatzel et al., 2004). CH4nd CO2 production rates will be affected by temperature in situherefore the effect of the stable laboratory conditions on CO2roduction rates should not be ignored. The size of the microbialiomass and the quality of the substrate become the main deter-inants of CO2 and CH4 production under constant temperature

nd moisture (Basiliko et al., 2007). CO2 and CH4 production maye connected in the way that CO2 as a product of organic matterecomposition may serve as a substrate for methanogenic microor-anisms producing methane. However, the relationship betweennaerobic CH4 and CO2 production rate is more likely to be a func-ion of overall microbial activity, rather than the use of CO2 as aubstrate for methanogenesis, as mentioned by Moore and Dalva1997).

.3. Effect of re-wetting on microbial methanogenic archaealommunity

The anaerobic CH4 production rate had similar range to thosen other studies (Moore and Dalva, 1993, 1997; Glatzel et al., 2004;asiliko et al., 2007). It is not clear whether variability in the rates ofH4 production is caused by environmental factors (e.g. substratevailability) or different numbers of diversity of methanogensommunities. The current view is that emissions of CH4 from north-rn peatlands vary as a function of temperature, water level, pH,egree of CH4 oxidation, vegetation structure (Bubier and Moore,994), however substrate and nutrient availability were observeds the most important factors influencing CH4 production rate inany studies (Glatzel et al., 2004; Basiliko et al., 2007). In our

tudy, higher CH4 production rates were measured for both intactites (minerotrophic mire and bog) as compared to the drainedites. The active population of methanogens was indicated by noag-time in CH4 production in peat samples from intact sites. Inontrast, a lag-time was observed before CH4 production in sam-les from drained sites. This lag-time indicated the presence, but

ow activity, of methanogens, as it was described by Basiliko et al.2003). The upward trend of the CH4 production rate was observeduring the whole period of incubation. This trend reflected the

ncreasing number and activity of methanogens; however no rela-ionship between number of methanogens and production rate wasbserved. This can be explained by the applied analysis for numberf methanogens. We used DNA analysis, which detected total num-

er of methanogens but not their activity, which can be detectednly by RNA analysis. A large part of the methanogens commu-ity can be present in soil in the inactive state due to limitation byany factors (Basiliko et al., 2003; Yavitt et al., 2005) and there-

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Z. Urbanová et al. / Ecological

ore a direct relation between number of methanogens and CH4roduction may not be observed.

However, our results showed that there is a potential forncreased CH4 production in drained peatlands after their re-

etting. Within a short period (three months) after re-wetting,he number of methanogens increased by one or two ordersf magnitude and the rate of methanogenesis increased by twor more orders of magnitude. The numbers of methanogensound in our samples correspond with those found by Kimt al. (2008). The methanogens were probably not limited byrganic substrate in soil samples from drained peatlands ashe rate of methanogenesis was increasing during the wholexperiment.

The community of methanogens from bog sites differed sig-ificantly from the communities from minerotrophic mires. Alsoethanogen diversity was lower at the bog sites compared to

he minerotrophic ones. Such differences were also observed bythers (Galand et al., 2005). Variability between communities ofethanogens was lower within bogs (BOG, BOGdr1) compared

o that within fens (FEN, FENdr1, FENdr2). These results probablyeflect the degradation state of drained sites and again confirmedhe conception of faster degradation of minerotrophic mires afterrainage as compared to bogs. This shift in methanogen commu-ity can be explained by changes in vegetation structure, which

s considered to be one of the most important factors influenc-ng archaeal methanogenic community composition (Galand et al.,005; Rooney-Varga et al., 2007). Also other environmental factorsuch as pH, substrate quality, temperature and water saturationan influence archaeal community composition and their activityBasiliko et al., 2003; Yavitt et al., 2005).

. Conclusions

After re-wetting of soils sampled from drained peatlands (simu-ated by anaerobic incubation of soils under laboratory conditions)i) phosphorus concentration (SRP) did not change in any soil,ii) concentration of ammonium and dissolved organic nitrogenDON) increased, but only in the drained fen, (iii) DOC increasedignificantly in the drained fen and degraded drained bog, (iv) CO2roduction decreased, (v) methane production and the number ofethanogens increased in all soils, and (vi) archaeal methanogenic

ommunity composition was also affected by re-wetting; it dif-ered significantly between the drained and pristine fens whereast was more similar between the drained and pristine bogs.

Overall, soils from minerotrophic mires reacted more dynam-cally to re-wetting as compared to bogs, and therefore, someutrients (especially nitrogen) and DOC leaching may be expected

rom drained fens after their water regime restoration. However,f compared to the state before restoration, ammonium and phos-horus leaching should not increase and leaching of nitrates andON should even decrease after restoration, especially during theegetation season. Further, CO2 production in soils of fens and bogshould decrease after their water regime restoration whereas CH4roduction in soils should increase. We cannot make any clearonclusions about CH4 emissions from the ecosystems as theyepend strongly on environmental factors and the actual activityf methanotrophs in situ.

cknowledgements

This study was supported by Projects No. 526/09/1545 of therant Agency of the Czech Republic, No. SP2d1/113/07 of the Czechinistry of Environment and MSM 600 766 5801. We thank the

umava National Park and Protected Landscape Area Administra-

M

eering 37 (2011) 1017– 1026 1025

ion, namely to Iva Bufková, for their support. We thank Dr. K.R.dwards for language correction, our technician T. Ríhová and ourtudents E. Jánská and J. Baxová for their perfect work in the labo-atory.

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