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Combined effects of carbon, nitrogen and phosphorus on CH4 production and denitrification in
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wetland sediments
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Sang Yoon Kim1, Annelies J. Veraart1, Marion Meima-Franke1 and Paul L.E. Bodelier1*
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1Netherlands
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Netherlands
Institute of Ecology (NIOO-KNAW), department of microbial ecology, Wageningen, the
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*Corresponding author: Paul L.E. Bodelier
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Phone: +31 (0)317473485
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Fax: +31 (0)317473675
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E-mail address: [email protected]
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Paper type: Regular article
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1. Introduction
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Methane (CH4) is the second most potent greenhouse gas in the atmosphere after carbon
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dioxide (CO2) and has 34 times the global-warming potential of CO2 over a 100-year horizon (IPCC,
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2013). Global atmospheric CH4 concentration has increased from pre-industrial level of 0.715 ppm to
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1.824 ppm in 2013 (WMO, 2014). Wetlands, including rice paddies, are the largest natural source of
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CH4 to the atmosphere, accounting for approximately 139 – 343 Tg CH4 yr-1. These ecosystems
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contribute 32 – 47% to the total global CH4 emissions (Denman, 2007). CH4 is produced by a complex
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microbial group which degrades organic matter by anaerobic methanogenesis (Conrad et al., 2007).
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Methanogenesis is mediated by acetoclastic, hydrogenotrophic, and methylotrophic methanogens that
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belong to the Euryarchaeota (Liu and Whitman, 2008).
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Methanogenesis can be regulated by various factors including temperature (Conrad, 2002;
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Glissman et al., 2004; Inglett and Inglett, 2013), pH (Wang et al., 1993; Ye et al., 2012), substrate
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availability (O'Connor et al., 2010), and availability of electron acceptors (D'Angelo and Reddy, 1999).
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Although a number of studies have been carried out to identify the main factors which control CH4
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dynamics from wetlands, the effect of nutrients on CH4 dynamics is poorly understood. Many studies
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point at nitrogen (N) as an important variable influencing CH4 cycles in wetland ecosystems (Bodelier,
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2011). Effects of N addition can act directly on the methanogenic community, but may have indirect
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effects, by stimulating the bacteria capable of denitrification. Denitrifiers and methanogens compete for
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organic carbon (C), and furthermore denitrification produces toxic intermediates (NO, NO 2, and N2O)
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which can inhibit methanogenic archaea in wetland sediments, thereby reducing CH4 production (Roy
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and Conrad, 1999). Although denitrification mitigates eutrophication effects in wetlands by reducing
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availability of reactive nitrogen, incomplete denitrification also contributes to the emission of N2O, which
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is an even more potent greenhouse gas than CH 4, with 300 times the global warming potential of CO 2
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(IPCC, 2013).
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Besides N, also phosphorus (P) can be an important regulating factor of methanogenesis. For
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example, in Dutch drainage ditches, the water column PO43- concentration was found to be an important
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predictor of CH4 emissions (Schrier-Uijl et al., 2011). Not only is P one of the most important nutrients
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influencing the microbial activity, including decomposition processes (Cleveland et al., 2002), but higher
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P concentrations also elevate microbial growth rates (Makino et al., 2003; Sterner and Elser, 2002).
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Recently, Medvedeff et al. (2014) suggested that P addition stimulated methanogenic activity in P-limited
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calcareous subtropical wetland soil, increasing the activity of methanogens directly or indirectly via
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fermentative bacteria that produce methanogenic substrates.
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Anthropogenic impacts on natural ecosystems have been increasing and largely influencing the
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balance of essential nutrients such as C, N and P for decades (Howarth and Marino, 2006; Wang et al.,
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2014). These changes in nutrient availability can impact the growth and activity of methanogens and
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denitrifiers, and may significantly influence CH4 cycling. However, little is known about the combined
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effects of C, N and P on CH4 cycling, and their impact on the interaction between methanogenesis and
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denitrification remains unclear in wetland ecosystems.
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The aim of this study was to investigate effects and possible interactions of N and P additions
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on CH4 production and denitrification in wetland sediments. More specifically we tested the following
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hypotheses: (i) sole P addition stimulates CH4 production due to elevated microbial growth rates
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including methanogens in wetland sediments, (ii) combined N and P addition enhance denitrification
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rates which lead to increased substrate competition between methanogens and denitrifiers, repressing
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CH4 production from wetland sediments. To this end incubation studies were performed with sediment
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derived from agricultural ditches in the Netherlands following methane production, denitrification,
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functional gene abundance and sediment physico-chemistry. Because effects of N and P are expected
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to act through competition for carbon, experiments were carried out with and without added C.
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2. Materials and methods
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2.1. Preparation and incubation of sediment slurries
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The sediment samples were collected from the top layer (0-5 cm) of a drainage ditch sediment
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in the spring of 2014 (Nigtevecht, The Netherlands; coordinates: 52° 16’ 41.92” N, 05° 01’ 40.11” E).
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The drainage ditches are common water management practices in the Netherlands. These ditches serve
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as important waterways and efficiently provide water to agricultural fields. The sediment pH was neutral
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(6.92) and had comparatively high organic C content (63 g kg-1) as well as CH4 production potential
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(106.9 nmol g-1 d.w hr-1) with low nutrient contents in pore water (Table 1). The sediment and pore water
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characteristics are presented in Table 1. Prior to use, sediment samples were passed through a
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stainless steel sieve (2 mm) by a wet-sieving method, and then stored at 4 °C until further processing.
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Anaerobic slurries were prepared (10 g fresh sediment and 13 ml amendment solution) in 150
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ml serum bottles. The bottles were vigorously shaken by hand to homogenize the sediment slurries,
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capped with sterilized butyl stoppers and then flushed with N2 for 1 hr. All bottles were incubated at
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25 °C in the dark for 12 days on a shaker (120 rpm). In order to verify potential interacting effects of N,
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P and C on CH4 production, we used a factorial design consisting of two C treatments (final conc. 0 or
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10 mM C as CH3COOH), three P treatments (final conc. 0, 1, and 10 mM P as KH2PO4) and 3 N
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treatments (final conc. 0, 2.5, and 5.0 mM N as KNO3), leading to a total of 18 combinations of C, N and
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P. Each experiment was carried out in triplicate.The final concentrations of C, N and P, are presented
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in Figure S1. N application level (0 - 5.0 mM, mean 2.5 mM) was chosen to resemble agriculture fields
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that receive high N loads (< 10 mM as 200 kg N ha-1), as suggested by Roy and Conrad (1999). In
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addition, P application level was determined by N to P ratios (0.25 - 5.0) based on possible N and P
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input from fertilization by chemical fertilizer and manure applications which globally ranged from 4.3 to
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5.7 (Mean ca. 5) by estimating N and P balance for a century from 1950 to 2050 (Bouwman et al., 2013).
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C addition (10 mM) was chosen to achieve a ratio of 4 C: 1 N, which provides sufficient C to the
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denitrifiers (Payne, 1981).
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Note that although added nutrients were in the high range of drainage ditch water column
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concentrations (Veraart, 2012), concentrations in the pore water, which are more relevant for this study,
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frequently reach the mM range (Veraart et al., 2015), with our highest added concentrations reflecting
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worst-case scenarios. Furthermore, nutrient concentrations are expected to increase with fertilizer
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application and heavy agricultural runoff in wetland ecosystem after extreme weather events, such as
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heavy rainfall, which due to global change is expected to occur more frequently in Western Europe.
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A control set of bottles (n = 3) was prepared to monitor the changes in headspace CH4
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concentration during incubation and an additional control series (n = 3) was used for chemical analysis
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of the pore water.
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2.2. Measurements of CH4 production and chemical parameters
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Methane concentrations were measured every day. Before sampling the headspace and pore
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water, bottles were shaken vigorously to homogenize the sediment slurries. Gas samples (200 µl) were
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taken using a gas-tight pressure-lock syringe flushed with N2. The CH4 in the headspace was measured
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using an Ultra GC gas chromatograph (Interscience, the Netherlands) equipped with Rt-Q-Bond (30 m,
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0.32 mm, ID) capillary column and a flame ionization detector (FID). The temperature of the column,
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injector and detector were adjusted to 80, 150, and 250 °C, respectively. Helium and H2 were used as
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the carrier and burning gases, respectively. Pore water samples (1 ml) were taken with sterile disposable
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syringes equipped with long needles flushed with N2 to prevent O2 leakage that might influence N
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mineralization processes such as ammonification and nitrification in the system. The pore water was
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transferred to Eppendorf tubes (2 ml) and centrifuged for 15 min at 15,000 × g at 4 °C. The supernatant
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was collected and stored at −20 °C until further analysis. Nutrient contents (NH4+, NO3- and NO2-) in the
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sediment were determined using an auto analyzer (QuAAtro, Seal analytical Inc., Beun de Ronde,
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Abcoude, The Netherlands). The pH in the slurries was measured directly after addition of the
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amendment solutions and after incubation.
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2.3. Denitrification potential
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After the methanogenic incubations, denitrification potential measurements were conducted
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using the acetylene inhibition method (Philippot et al., 2013; Qin et al., 2012). Briefly, 5g of fresh slurry
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from the bottles after incubation was transferred to new bottles (150 ml) and 7.5 ml of a solution
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containing KNO3 (1 mM) and glucose (1 mM) was added. The bottles were sealed and capped with
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sterilized butyl stoppers and then flushed with N2 for 10 min. Acetylene was added (10% v/v of the
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headspace) and the bottles were incubated at 25 °C on a shaker (120 rpm). Nitrous oxide (N2O) in the
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headspace was measured after 24 hrs. of incubation using a TRACE 1300 (Thermo Scientific, the
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Netherlands) equipped with HS-Q (1m, 2.0 mm, ID) packed column fitted with an electron capture
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detector (ECD). The temperature of the column and injector was 90 °C and the detector was set at
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350 °C. N2 gas was used as the carrier.
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2.4. Quantification of mcrA and nirS gene copy numbers in sediment slurries
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After incubation, DNA was extracted from 0.10 g of freeze-dried sediment following a slightly
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modified protocol based on the FastDNA spin kit for soil (MP Biomedicals, Solon, OH) previously
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described in detail in Wang et al. (2012). Nucleic acids were routinely quantified using a NanoDrop 1000
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spectrophotometer (Thermo) before quantitative PCR (qPCR). The copy numbers of the mcrA gene,
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encoding the methyl coenzyme-M reductase were used as proxy for methanogenic abundance and the
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copy numbers of the nirS gene, encoding the cytochrome cd1 nitrite reductase were used as proxy for
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denitrifier abundance. Primer sets of mlas/mcrA-rev were used for mcrA (Steinberg and Regan, 2008)
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and nirScd3af/nirSR3cd for nirS (Throbäck et al., 2004). Real-time qPCR was performed in a Rotor-
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Gene Q real-time PCR cycler (Qiagen, the Netherlands). Briefly, qPCR reaction (total volume 20 µl) for
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mcrA gene contained 10 µl of reaction mixtures 2X SensiFAST SYBR (BIOLINE, the Netherlands), 3.5
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µl of forward primers (4 pmol µl-1), 3.5 µl reverse primers (5 pmol µl-1), 1 µl Bovine Serum Albumin (5
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mg ml-1; Invitrogen, the Netherlands), and 2 µl diluted template DNA (2 ng µl-1). Amplification was carried
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out as follows: for the mcrA gene, initial denaturation at 95 °C for 3 min, followed by 45 cycles of
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denaturation at 95 °C for 10 sec, annealing at 60 C for 10 sec, and extension at 72 °C for 25 sec. qPCR
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reaction (total volume 20 µl) for nirS gene contained 10 µl of reaction mixtures 2X SensiFAST SYBR
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(BIOLINE, the Netherlands), 2 µl of forward primers (5 pmol µl-1), 2 µl reverse primers (5 pmol µl-1), 1 µl
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Bovine Serum Albumin (5 mg ml-1; Invitrogen, the Netherlands), and 2.5 µl diluted template DNA (2 ng
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µl-1). Amplification was carried out as follows: for the nirS gene, initial denaturation at 95 °C for 3 min,
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followed by 40 cycles of denaturation at 95 °C for 10 sec, annealing at 60 °C for 10 sec, and extension
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at 72 °C for 20 sec and 86 °C for 5 sec. In each qPCR, the fluorescence signal was obtained at 72 °C
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after each cycle, and melt curves were obtained from 70 °C to 99 °C (1 °C temperature by rising).
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Amplicon specificity was determined from the melt curve. To avoid the inhibitory effects of substances
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co-extracted with the DNA, amplification of serial dilutions was carried out for the slurry samples in each
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treatment.
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2.5. Statistical analysis
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Statistical analyses were conducted using R studio software (ver.2.6.0). Determination of
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differences between parameters was performed through three-way analysis of variance (ANOVA)
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including N, P and C additions and their interactions.
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3. Results
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3.1. Effects of C, N and P on CH4 production and potential denitrification
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3.1.1 CH4 production and its lag phase
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C, N, P and their interaction all significantly affected CH 4 production and its lag phase (Figures
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1 and 2, Table 2, three-way ANOVA, see Table 3). Overall, we observed most CH4 in the headspace of
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samples where no N was added, of which samples with added C had the highest final CH4
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concentrations (113.2 µM CH4). When C was added, effects of N and P clearly interacted: the lowest
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CH4 production (1.14 µM CH4) was found in the samples where the highest P (10 mM) and N (5 mM)
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were added. Combined C and P addition at 10 mM increased the lag phase of CH4 production, especially
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in combination with N addition (Figure 1 and Table 2).
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3.1.2. Denitrification potential
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Denitrification potential was on average 28.2 nmol N g-1 d.w-1, and showed interacting effects of
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C, N and P (Figure 2, three-way ANOVA, see Table 3). When no P was added, addition of N and C
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moderately stimulated denitrification. With 1 mM P added, C addition moderately enhanced
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denitrification, but for the treatment without C the effect of N is not clear. Interestingly, a very different
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pattern can be seen for the highest P addition (10 mM), where there seems to be an additive inhibitory
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effect of N and C addition: the lowest denitrification potential (8.28 nmol N g-1 d.w-1) was found when P,
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N and C were added.
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3.2. Changes in biogeochemical properties
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3.2.1. Changes of NO3-, NO2-, and NH4+ in pore water
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NO3- was consumed within 48 hr in sediment slurries without P addition, but decreased slower
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in slurries where P was added, irrespective of C addition. In sediment slurries not amended with N and
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C, no NO2- was detected during the incubation period. NO2- was only observed in the sediment to which
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both C and N were added, and was higher in the slurries with added P (10 mM). For example, there was
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a NO2- accumulation, up to 0.27 mM in the slurries where C, N and P was added (C, N, and P: 10, 5.0,
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and 10 mM). The NO2- peak occurs when NO3- levels are still decreasing, and vanishes with
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disappearing NO3-. Interestingly, although without P addition, only small fluctuations in NH 4+ occurred in
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all treatments, when C, N and P were added in combination, an increase in NH 4+ above the baseline
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variation can be observed (especially at N 2.5 mM), which later decreases again. Overall, NH 4+
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concentrations were maintained at a higher level in P-added slurries (0.036 - 0.103 mM on average)
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than in slurries where no P was added (0.020 - 0.034 mM on average) (Figure 3).
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3.2.2. Changes of pH
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C addition decreased the pH of the sediment slurries, but during incubation pH in these slurries
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neutralized again (Figure S2). P addition also acidified the slurries, with lower pH in P10 slurries (ca.
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6.30) than in P0 and P1 slurries (ca. 6.73), and these effects remained after incubation.
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3.2.3. mcrA and nirS gene copy numbers
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The abundances of the mcrA gene and nirS gene were determined by qPCR after slurry
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incubation (Figure 4). The mcrA gene ranged from 6.78 × 107 to 1.10 × 108 and nirS gene copy numbers
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ranged from 5.21 × 106 to 6.62 × 107 copies g-1 dry sediment. Unexpectedly, the copy numbers of mcrA
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genes were higher in samples without C addition but increased with P addition except N0 treatment.
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The effect of N addition on mcrA gene copy was not observed under C addition. In particular, N addition
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gradually increased nirS gene copy numbers in slurries with added P or C, but copy numbers decreased
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when P and C were both added. We found a weak correlation but significant negative exponential
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relations between CH4 production and mcrA gene copy numbers (R2=0.164, P=0.051). A positive
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exponential correlation between nirS gene copy numbers and denitrification potential (R2=0.288,
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P<0.01) was also observed.
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3.3. Relationship between CH4 production and denitrification potential
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P additions (P1 and P10) without added N showed similar patterns as the control without added
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P, and did not influence CH4 production and denitrification potential, irrespective of C addition (Figure
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5). N additions without added P decreased CH4 production but increased denitrification potential in
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sediments with and without added C. Combined N and P addition showed additive effects on CH4
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production and denitrification potential. However, the processes had a different response to C addition:
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when C was added, CH4 production and denitrification were more strongly inhibited at high N:P ratios
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(2.5 - 5.0) than at low N:P ratios (0.25 - 0.5).
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4. Discussion
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We hypothesized that sole and combined effects of N and P on methanogenesis and
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denitrification in wetlands sediments act through competition for C. Experiments were carried out with
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and without added C to measure CH4 production potential and potential denitrification as well as
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biogeochemical properties during slurry incubations. The present study shows combined effects of N
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and P on CH4 and N cycles in wetland sediments, and indicates that these effects may indeed occur
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through competition for C - albeit in a more complex interaction than initially hypothesized - which will
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be discussed in the following sections.
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4.1. Effects of C, N and P on CH4 production and denitrification
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Our study showed that CH4 production was significantly affected by C addition in sediment
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slurries. It is well-known that acetate can be directly utilized by acetogenic methanogens to produce CH 4
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under anaerobic conditions (Conrad, 2007). This stimulation was not reflected in the mcrA gene
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abundance, for which we have no definite explanation other than referring to other studies which also
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found no relation between gene abundance and methane production (Cadillo‐Quiroz et al., 2006;
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Galand et al., 2003). Possibly, acetate and phosphate have formed acetylphosphate, which may have
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affected enzyme functioning in the metabolic pathways and qPCR process.
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N additions inhibited CH4 production (Figure 2). Its effect increased with increasing N
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concentration. The inhibitory effect of N compounds on the methanogenic microbial community can be
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caused by denitrification intermediates (NO, NO2-) that are toxic for methanogenic archaea, or by
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competition between denitrifiers and methanogens for acetate (Kluber and Conrad, 1998; Roy and
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Conrad, 1999). NO2- was only produced in excess when C and N were added, and this effect was much
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stronger under high P (Figure 3). This result indicates that the increased C/N ratio influenced the first
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denitrification steps and that the N:P ratio can also affect NO 2- accumulation. The accumulation of NO2-
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stops when NO3- levels decrease, which is also expected from theoretical denitrification energetics (van
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de Leemput et al., 2011).
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Roy and Conrad (1999) speculated that competition for substrate between denitrifiers and
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methanogens is not the main mechanism of suppression of methanogenesis in rice soil. However, we
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observed inhibitory effects of N addition on CH4 production and lag phase in slurries where no C was
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added (Figure 2). In these slurries, no NO2- accumulation was observed (Figure 3), which indicates that
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the inhibitory effect is mainly due to the competition for C between methanogens and other
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microorganisms.
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Our study showed that CH4 production was not directly influenced by P addition in sediment
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slurries, irrespective of C addition. However, P addition slowed down NO 3- uptake by denitrifiers, which
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can increase C availability for methanogens by decreasing competition. Therefore, we expected CH 4
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production to be increased during the delayed NO 3- uptake. Different to our expectation, CH4 production
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was significantly lower when N and P were added, irrespective of C addition (Figure 2). It appears that
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low N:P ratios decreased denitrification. A negative correlation between P and denitrification was also
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observed in the sediment of C-poor shallow lakes (Veraart, 2012). On the other hand, seasonal positive
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correlations between total P and denitrification were observed in a meta-analysis of denitrification across
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aquatic ecosystems (Piña-Ochoa and Álvarez-Cobelas, 2006). However, P-release from the sediment
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and denitrification occur under similar anoxic conditions, confounding ecosystem observations, and
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making negative correlations between P and denitrification even more interesting. The potential
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inhibitory effect of P on denitrification may be direct or indirect. Potential direct effects are not well
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understood, but may be through P-sensitivity of the denitrifiers due to adaptation to the low prevailing P
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conditions in this sediment. Indirect effects through enhanced competition for C and N with other
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bacterial populations are also possible. We used copy numbers of nirS, reflecting the denitrifier
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community containing the cytochrome cd1 nitrite reductase, as a proxy for denitrifier abundance. This
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may give an incomplete picture, because part of the denitrifiers will use nirK rather than nirS, and mere
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gene presence does not directly reflect functional activity. Nonetheless abundance of nirS positively
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correlated with potential denitrification (Figure S3). Copy numbers of nirS were also highest when both
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P and N were present, which may hint at an N and P co-limitation of the denitrifying community. However,
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much like mcrA copy numbers, this effect of N and P on nirS copy numbers was only observed when
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no C was added.
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4.2. Competition between denitrification, DNRA and methanogenesis
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C addition may have shifted the competitive balance of different nitrate respiring and C
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degrading communities. Possibly, communities carrying out DNRA (dissimilatory nitrate reduction to
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ammonium) may have been better competitors for acetate under N and P addition, suppressing the
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abundance of nirS-denitrifiers as well as methanogenic activity. DNRA-communities compete with
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denitrifiers for C and N, as these pathways occur under similar conditions (Francis et al., 2007; Scott et
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al., 2008; Tiedje, 1988; Yin et al., 2002). The C:N ratio is a key determinant of the competitive outcome
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between both processes, favoring DNRA at high C availability (Smith, 1982), in line with our findings of
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lower denitrification under combined N, P and C addition. Our ‘DNRA-competition’-hypothesis is further
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supported by the observed increase in NH 4+ concentrations in slurries with combined N, P and
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application (Figure 2). However, we estimate that about 10% of the NO3- was converted to NH4+ - rather
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than gaseous nitrogen - in these slurries, leaving part of the difference in denitrification potential
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unexplained. A closed mass balance or isotope tracer approach would be optimal to accurately trace N-
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conversions.
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Interestingly, combined C and P addition seems to enhance initial ammonification and DNRA,
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suggesting that these steps of the N cycle are P limited in this sediment. Clearly, further studies are
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needed to test P-limitation of ammonification, DNRA, and denitrification and consequently its
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implications for CH4 cycling. Isotope tracing methods based on single cells (Krause et al., 2014) such
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as stable isotope tracers and nano-SIMS, in combination with bacterial community analyses to identify
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shifts between dominant populations and N-respiring pathways (Kraft et al., 2014) will be promising
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ways to study this.
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5. Conclusions
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We have summarized the main findings of this study in a conceptual scheme depicted in Figure
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6. The effects of N, P and C additions and the interaction between methanogenesis and denitrification
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turned out to be more complicated than hypothesized. Instead of simply stimulating growth of
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methanogens or denitrifiers by N or P additions we observed an interactive effect of C, N and P which
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may be explained by a modulation of electron flow towards DNRA, consuming electron donors in the
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process. Despite the fact that this hypothesis still has to be verified experimentally it is safe to conclude
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that P might play an important modulating role in carbon degradation and C-N cycle interactions in
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wetland sediments with possible consequences for greenhouse gas emissions from these ecosystems.
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Acknowledgements
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This work was supported a grant (823.001.008) of the Netherlands Organisation for Scientific Research
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(NWO). This work was supported by the National Research Foundation of Korea (NRF) grant funded by
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the Korea government (MSIP) (No. NRF-2013R1A2A2A07068946). We thank Anne Steenbergh for help
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with the statistical analysis. This publication is publication no. xxxx of the Netherlands Institute of
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Ecology.
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Table 1. Characteristics of the sediment and pore water before the experiment.
Parameters
Value
Sediment
pH
6.92
Total C (g kg-1)
63.0
Total N (g kg-1)
5.42
C:N ratio
11.6
Potential CH4 production (nmol g-1 d.w hr-1)
106.9
Pore water
Electrical conductivity (µS cm-1)
853
Dissolved organic C (mg L-1)
14.0
Dissolved inorganic C (mg L-1)
3.8
NH4+ (mg L-1)
0.43
PO43- (mg L-1)
2.0
N:P ratio
0.51
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Table 2. Variables of methanogenic potentials in sediment slurries of different nutrient addition
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treatments.
Treatment
Variables
Without
P0
C addition
P1
P10
With C addition
P0
P1
P10
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* Initiation
315
regressions).
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** BD
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*** R2 values
Lag phase (hr)*
CH4 production rate (umol g-1 hr-1)
R2***
N0
39.5 ± 6.80
0.040 ± 0.003
0.996
N2.5
48.2 ± 6.40
0.006 ± 0.001
0.988
N5.0
52.2 ± 31.8
0.006 ± 0.002
0.978
N0
23.4 ± 10.1
0.045 ± 0.001
0.999
N2.5
74.5 ± 32.0
0.005 ± 0.001
0.978
N5.0
81.6 ± 26.6
0.005 ± 0.002
0.959
N0
12.7 ± 0.90
0.038 ± 0.002
0.998
N2.5
-
BD**
-
N5.0
-
BD
-
N0
23.3 ± 1.20
0.344 ± 0.004
0.994
N2.5
38.0 ± 1.00
0.102 ± 0.003
0.907
N5.0
51.7 ± 0.60
0.087 ± 0.004
0.980
N0
26.0 ± 1.30
0.327 ± 0.007
0.992
N2.5
57.3 ± 0.20
0.108 ± 0.002
0.946
N5.0
74.2 ± 0.70
0.091 ± 0.004
0.969
N0
34.8 ± 2.50
0.126 ± 0.041
0.942
N2.5
109.0 ± 3.900
0.025 ± 0.010
0.951
N5.0
-
BD
-
of linear phase of CH4 production in the data using four time points (extrapolated from linear
means below detection limits (CH4 production rate < 0.001).
were estimated from CH4 production in the data.
14 | P a g e
318
Table 3. Three-way ANOVA results. All variables except pH were log-transformed to improve normality.
variable
Total CH4
CH4 lag-phase
production
319
Potential
CH4 production
denitrification rate
rate
mcrA
nirS
pH
pH
(before incubation)
(after incubation)
F
P
F
P
F
P
F
P
F
P
F
P
F
P
F
P
CxNxP
19.91
<0.001
6.392
0.016
108.33
<0.001
0.646
0.427
6.341
0.018
25.164
<0.001
0.119
0.7317
6.69
0.013
NxP
43.25
<0.001
11.638
0.002
85.60
<0.001
8.736
0.006
0.407
0.529
5.936
0.022
6.993
0.011
0.237
0.6289
CxP
15.60
<0.001
42.418
<0.001
327.31
<0.001
20.37
<0.001
0.501
0.485
52.418
<0.001
2.860
0.098
0.024
0.877
CxN
1.08
0.305
0.173
0.680
39.93
<0.001
13.88
<0.001
0.017
0.898
3.769
0.064
0.038
0.564
39.56
<0.001
C
258.12
<0.001
10.354
0.003
31.35
<0.001
294.8
<0.001
4.046
0.054
28.208
<0.001
693.7
<0.001
5.392
0.0247
N
250.82
<0.001
63.27
0.001
0.01
0.909
117.7
<0.001
1.540
0.225
10.033
0.004
38.495
<0.001
93.71
<0.001
P
63.22
<0.001
1.715
0.199
170.61
<0.001
29.47
<0.001
0.105
0.748
20.247
<0.001
838.7
<0.001
2404.18
<0.001
Note) F value degrees of freedom = F1; P< 0.05, P< 0.01, and P< 0.001 denotes significance at the 5, 1, and 0.1 % levels, respectively.
15 | P a g e
320
Figure legends
321
Figure 1. Changes in CH4 concentrations in different nutrient-amended slurries during incubation with
322
C or without C added. Bars represent standard deviations (n=3). Each panel shows a different P addition
323
level and the lines per panel show different N addition levels.
324
Figure 2. Total CH4 production and denitrification potential after incubation. Bars represent standard
325
deviations (n=3).
326
Figure 3. Changes in concentrations of nitrate (NO3-), nitrite (NO2-), and ammonium (NH4+) in pore water
327
in different nutrient-amended slurries during incubation with C or without C addition. Bars represent
328
standard deviations (n=3).
329
Figure 4. Abundances of the mcrA and nirS genes in different nutrient-amended sediments after
330
incubation. Bars represent standard deviations (n=3).
331
Figure 5. Effect of N:P ratio on CH4 production and denitrification.
332
Figure 6. Conceptual scheme of observed and hypothetical effects of the applied treatments in this
333
study.
16 | P a g e
Without C addition (0 mM)
N0
25
With C addition (10 mM)
P0
P0
N0
N2.5
N2.5
N5.0
N5.0
120
100
20
80
15
60
10
40
5
20
0
0
50
100
CH4 in headspace (umol)
25
150
P1
200
250
0
50
100
X Data
150
P1
200
250
120
X Data
100
20
80
15
60
10
40
5
CH4 in headspace (umol)
0
20
0
0
0
50
100
25
150
P10
200
250
0
50
100
X Data
150
P10
200
250
120
X Data
100
20
80
15
60
10
40
5
20
0
0
0
334
50
100
150
200
Incubation time (hr)
250
0
50
100
150
200
250
Incubation time (hr)
335
336
337
Figure 1.
17 | P a g e
without C addition
with C addition (10 mM)
60
-1
Produced CH4 (umol g d.w)
50
40
30
20
10
1
N0
N2.5
N5.0
N0
N2.5
N5.0
N0
N2.5
N5.0
-1
Denitrification potential (nmol N2O g d.w hr )
0
P10
P1
(mM)
P0
40
-1
Treatment
30
20
10
0
N0
N2.5
P0
N5.0
N0
N2.5
N5.0
P1
(mM)
N0
N2.5
N5.0
P10
Treatment
338
339
340
341
Figure 2.
18 | P a g e
with C addition (P10)
N 0 mM
N 2.5 mM
N 5.0 mM
1,2
0,9
0,9
0,9
0,6
0,6
0,3
0,3
Y Data
1,2
0,6
0,3
0,0
0,0
0,0
0
50
100
0,5
150
200
250
0
50
100
150
200
250
0
50
100
0,5
X Data
150
200
250
0
50
100
150
200
X Data
N 0 mM
250
N 2.5 mM
N 5.0 mM
0,5
0,4
0,4
0,3
0,3
0,2
0,2
0,2
0,1
0,1
0,1
0,0
0,0
Y Data
0,4
1,2
0,3
0
50
100
150
200
250
0,0
0
50
100
X Data
150
200
250
0
50
100
150
200
250
0
50
100
150
200
X Data
N 0 mM
250
N 2.5 mM
N 5.0 mM
0,2
0,2
0,2
0,1
0,1
0,1
0,0
0,0
0,0
0
50
100
150
200
Incubation period (hr)
342
250
0
50
100
150
200
Incubation period (hr)
250
0
50
100
150
200
Incubation period (hr)
250
0
50
100
150
200
NO 3- concentration in pore water (mM)
without C addition (P10)
NO 2- concentration in pore water (mM)
with C addition (P0)
NH4+ concentration in pore water (mM)
NH4+ concentration in pore water (mM)
NO 2- concentration in pore water (mM)
NO 3- concentration in pore water (mM)
without C addition (P0)
250
Incubation period (hr)
343
344
Figure 3.
19 | P a g e
nirS gene
without C addition
with C addition (10 mM)
1,2e+9
8,0e+7
1,0e+9
6,0e+7
8,0e+8
4,0e+7
6,0e+8
4,0e+8
2,0e+7
2,0e+8
N0
N2.5
N5.0
N0
without P addition (P0)
Treatments
345
N2.5
N5.0
with P addition (P10)
N0
N2.5
N5.0
N0
without P addition (P0)
N2.5
nirS gene copies g-1 dried sediment
mcrA gene copies g-1 dried sediment
mcrA gene
N5.0
with P addition (P10)
Treatments
346
347
348
Figure 4.
20 | P a g e
9
50
40
30
6
20
10
3
-1
12
with C addition
CH4 production (umol g d.w)
No addition (control)
Sole N (2.5 mM)
Sole N (5.0 mM)
Sole P (1.0 mM)
Sole P (10 mM)
NP (N:P ratio 0.25)
NP (N:P ratio 0.50)
NP (N:P ratio 2.50)
NP (N:P ratio 5.00)
-1
CH4 production (umol g d.w)
without C addition
0
0
25
30
35
40
-1
45
-1
Denitrification potential (nmol N2O g d.w hr )
349
10
15
20
25
30
-1
35
-1
Denitrification potential (nmol N2O g d.w hr )
350
351
352
Figure 5.
21 | P a g e
353
354
355
356
Figure 6.
22 | P a g e
357
Supporting Information Legends
358
Figure S1. Experimental scheme for nutrient and substrate additions.
359
Figure S2. Initial and final pH of incubation in different nutrient-amended sediments. Bars represent
360
standard deviations (n=3).
361
Figure S3. Relationships between functional gene abundances and their activities.
23 | P a g e
P(0 mM)
C(0 mM )
P+
(1 mM)
P++
(10 mM)
•
N-
•
N+ (2.5 mM)
•
N++ (5.0 mM)
•
N-
•
N+ (2.5 mM)
•
N++ (5.0 mM)
•
N-
•
N+ (2.5 mM)
•
N++ (5.0 mM)
•
N-
•
N+ (2.5 mM)
•
N++ (5.0 mM)
•
N-
•
N+ (2.5 mM)
•
N++ (5.0 mM)
•
N-
•
N+ (2.5 mM)
•
N++ (5.0 mM)
(0 mM)
(0 mM)
(0 mM)
Wetland
sediment
P(0 mM)
C+
(10 mM)
P+
(1 mM)
P++
(10 mM)
(0 mM)
(0 mM)
(0 mM)
362
363
364
365
Figure S1.
24 | P a g e
without C addition
with C addition (10 mM)
pH (after additions)
7.0
6.5
6.0
5.5
N0
N2.5
N5.0
N0
P0
N2.5
N5.0
N0
P1
(mM)
7.0
N2.5
N5.0
P10
pH (after incubations)
Treatment
6.5
6.0
5.5
N0
N2.5
P0
N5.0
N0
N2.5
P1
N5.0
N0
N2.5
N5.0
P10
(mM)
Treatment
366
367
368
369
Figure S2.
25 | P a g e
nirS
-10
2
Y = 20.8-33.9*(1-exp(-9.0*10 X), R =0.164, p=0.051
40
30,0
30
20
20,0
10
0
10,0
Y = 20.4+13.2*(1-exp(-8.4*10-8X), R2=0.288, P<0.01
0,0
370
2,0e+8
4,0e+8
6,0e+8
8,0e+8
1,0e+9
mcrA gene copy numbers g-1 dried sediment
1,2e+9
0,0
2,0e+7
4,0e+7
-1
-1
50
Denitrification potential (nmol g d.w hr )
-1
-1
CH4 production potential (nmol g d.w hr )
mcrA
6,0e+7
nirS gene copy numbers g-1 dried sediment
371
372
373
Figure S3.
26 | P a g e
374
REFERENCES
375
Bodelier, P.L., 2011. Interactions between nitrogenous fertilizers and methane cycling in wetland and
376
377
upland soils. Current Opinion in Environmental Sustainability 3(5), 379-388.
Bouwman, L., Goldewijk, K.K., Van Der Hoek, K.W., Beusen, A.H.W., Van Vuuren, D.P., Willems, J.,
378
Rufino, M.C., Stehfest, E., 2013. Exploring global changes in nitrogen and phosphorus cycles
379
in agriculture induced by livestock production over the 1900–2050 period. Proceedings of the
380
National Academy of Sciences 110(52), 20882-20887.
381
Cadillo‐Quiroz, H., Bräuer, S., Yashiro, E., Sun, C., Yavitt, J., Zinder, S., 2006. Vertical profiles of
382
methanogenesis and methanogens in two contrasting acidic peatlands in central New York
383
State, USA. Environ Microbiol 8(8), 1428-1440.
384
Cleveland, C.C., Townsend, A.R., Schmidt, S.K., 2002. Phosphorus Limitation of Microbial Processes
385
in Moist Tropical Forests: Evidence from Short-term Laboratory Incubations and Field Studies.
386
Ecosystems 5(7), 0680-0691.
387
388
Conrad, R., 2002. Control of microbial methane production in wetland rice fields. Nutr Cycl
Agroecosys 64(1-2), 59-69.
389
Conrad, R., 2007. Microbial ecology of methanogens and methanotrophs. Adv Agron 96, 1-63.
390
Conrad, R., Chan, O., Claus, P., Casper, P., 2007. Characterization of methanogenic Archaea and
391
stable isotope fractionation during methane production in the profundal sediment of an
392
oligotrophic lake (Lake Stechlin, Germany). Limnol Oceanogr 52(4), 1393.
393
394
395
D'Angelo, E.M., Reddy, K.R., 1999. Regulators of heterotrophic microbial potentials in wetland soils.
Soil Biology and Biochemistry 31(6), 815-830.
Denman, K., 2007. Denman, K.L., G. Brasseur, A. Chidthaisong, P. Ciais, P.M. Cox, R.E. Dickinson,
396
D. Hauglustaine, C. Heinze, E. Holland, D. Jacob, U. Lohmann, S Ramachandran, P.L. da
397
Silva Dias, S.C. Wofsy and X. Zhang, 2007: Couplings Between Changes in the Climate
398
System and Biogeochemistry. In: Climate Change 2007: The Physical Science Basis.
399
Contribution of Working Group I to the Fourth Assessment Report of the Intergovernmental
400
Panel on Climate Change [Solomon, S., D. Qin, M. Manning, Z. Chen, M. Marquis, K.B.
401
Averyt, M.Tignor and H.L. Miller (eds.)]. Cambridge University Press, Cambridge, United
402
Kingdom and New York, NY, USA.
27 | P a g e
403
404
405
406
407
Francis, C.A., Beman, J.M., Kuypers, M.M.M., 2007. New processes and players in the nitrogen cycle:
the microbial ecology of anaerobic and archaeal ammonia oxidation. ISME J 1(1), 19-27.
Galand, P.E., Fritze, H., Yrjälä, K., 2003. Microsite‐dependent changes in methanogenic populations
in a boreal oligotrophic fen. Environ Microbiol 5(11), 1133-1143.
Glissman, K., Chin, K.-J., Casper, P., Conrad, R., 2004. Methanogenic pathway and archaeal
408
community structure in the sediment of eutrophic Lake Dagow: effect of temperature. Microb
409
Ecol 48(3), 389-399.
410
411
412
413
414
Howarth, R.W., Marino, R., 2006. Nitrogen as the limiting nutrient for eutrophication in coastal marine
ecosystems: evolving views over three decades. Limnol Oceanogr 51(1), 364-376.
Inglett, P.W., Inglett, K.S., 2013. Biogeochemical changes during early development of restored
calcareous wetland soils. Geoderma 192(0), 132-141.
IPCC, 2013. Climate Change 2013: The Physical Science Basis. Contribution of Working Group I to
415
the Fifth Assessment Report of the Intergovernmental Panel on Climate Change [Stocker,
416
T.F., D. Qin, G.-K. Plattner, M. Tignor, S.K. Allen, J. Boschung, A. Nauels, Y. Xia, V. Bex and
417
P.M. Midgley (eds.)]. Cambridge University Press, Cambridge, United Kingdom and New
418
York, NY, USA, 1535 pp.
419
420
Kluber, H.D., Conrad, R., 1998. Inhibitory effects of nitrate, nitrite, NO and N2O on methanogenesis by
Methanosarcina barkeri and Methanobacterium bryantii. Fems Microbiol Ecol 25(4), 331-339.
421
Kraft, B., Tegetmeyer, H.E., Sharma, R., Klotz, M.G., Ferdelman, T.G., Hettich, R.L., Geelhoed, J.S.,
422
Strous, M., 2014. The environmental controls that govern the end product of bacterial nitrate
423
respiration. Science 345(6197), 676-679.
424
Krause, S., Le Roux, X., Niklaus, P.A., Van Bodegom, P.M., Lennon, J.T., Bertilsson, S., Grossart, H.-
425
P., Philippot, L., Bodelier, P.L., 2014. Trait-based approaches for understanding microbial
426
biodiversity and ecosystem functioning. Frontiers in Microbiology 5, 251
427
210.3389/fmicb.2014.00251
428
429
430
Liu, Y., Whitman, W.B., 2008. Metabolic, Phylogenetic, and Ecological Diversity of the Methanogenic
Archaea. Annals of the New York Academy of Sciences 1125(1), 171-189.
Makino, W., Cotner, J.B., Sterner, R.W., Elser, J.J., 2003. Are bacteria more like plants or animals?
431
Growth rate and resource dependence of bacterial C : N : P stoichiometry. Funct Ecol 17(1),
432
121-130.
28 | P a g e
433
Medvedeff, C.A., Inglett, K.S., Inglett, P.W., 2014. Evaluation of direct and indirect phosphorus
434
limitation of methanogenic pathways in a calcareous subtropical wetland soil. Soil Biology and
435
Biochemistry 69(0), 343-345.
436
O'Connor, F.M., Boucher, O., Gedney, N., Jones, C.D., Folberth, G.A., Coppell, R., Friedlingstein, P.,
437
Collins, W.J., Chappellaz, J., Ridley, J., Johnson, C.E., 2010. Possible role of wetlands,
438
permafrost, and methane hydrates in the methane cycle under future climate change: A
439
review. Reviews of Geophysics 48(4), RG4005.
440
Payne, W.J., 1981. Denitrification. John Wiley & Sons Inc.
441
Philippot, L., Spor, A., Henault, C., Bru, D., Bizouard, F., Jones, C.M., Sarr, A., Maron, P.A., 2013.
442
443
444
445
446
447
Loss in microbial diversity affects nitrogen cycling in soil. Isme Journal 7(8), 1609-1619.
Piña-Ochoa, E., Álvarez-Cobelas, M., 2006. Denitrification in Aquatic Environments: A Cross-system
Analysis. Biogeochemistry 81(1), 111-130.
Qin, S.P., Hu, C.S., Oenema, O., 2012. Quantifying the underestimation of soil denitrification potential
as determined by the acetylene inhibition method. Soil Biol Biochem 47, 14-17.
Roy, R., Conrad, R., 1999. Effect of methanogenic precursors (acetate, hydrogen, propionate) on the
448
suppression of methane production by nitrate in anoxic rice field soil. Fems Microbiol Ecol
449
28(1), 49-61.
450
Schrier-Uijl, A., Veraart, A., Leffelaar, P., Berendse, F., Veenendaal, E., 2011. Release of CO2 and
451
CH4 from lakes and drainage ditches in temperate wetlands. Biogeochemistry 102(1-3), 265-
452
279.
453
Scott, J.T., McCarthy, M., Gardner, W., Doyle, R., 2008. Denitrification, dissimilatory nitrate reduction
454
to ammonium, and nitrogen fixation along a nitrate concentration gradient in a created
455
freshwater wetland. Biogeochemistry 87(1), 99-111.
456
457
458
Smith, M.S., 1982. Dissimilatory Reduction of NO2− to NH4+ and N2O by a Soil Citrobacter sp.
Applied and environmental Microbiology 43(4), 854-860.
Steinberg, L.M., Regan, J.M., 2008. Phylogenetic comparison of the methanogenic communities from
459
an acidic, oligotrophic fen and an anaerobic digester treating municipal wastewater sludge.
460
Applied and environmental microbiology 74(21), 6663-6671.
461
462
Sterner, R.W., Elser, J.J., 2002. Ecological stoichiometry: the biology of elements from molecules to
the biosphere. Princeton University Press.
29 | P a g e
463
Throbäck, I.N., Enwall, K., Jarvis, Å., Hallin, S., 2004. Reassessing PCR primers targeting nirS, nirK
464
and nosZ genes for community surveys of denitrifying bacteria with DGGE. Fems Microbiol
465
Ecol 49(3), 401-417.
466
467
468
Tiedje, J.M., 1988. Ecology of denitrification and dissimilatory nitrate reduction to ammonium. Biology
of anaerobic microorganisms 717, 179-244.
van de Leemput, I.A., Veraart, A.J., Dakos, V., de Klein, J.J., Strous, M., Scheffer, M., 2011.
469
Predicting microbial nitrogen pathways from basic principles. Environ Microbiol 13(6), 1477-
470
1487.
471
Veraart, A.J., 2012. Denitrification in Ditches, Streams and Shallow Lakes. Wageningen University.
472
Veraart, A.J., Steenbergh, A., Ho, A., Kim, S.Y., Bodelier, L.E.P., 2015. Beyond nitrogen: The
473
474
importance of phosphorus for CH4 oxidation in soils and sediments, Geoderma (in revision).
Wang, J., Krause, S.M.B., Muyzer, G., Meima-Franke, M., Laanbroek, H.J., Bodelier, P.l.E., 2012.
475
Spatial patterns of iron- and methane-oxidizing bacterial communities in an irregularly flooded,
476
riparian wetland. Frontiers in Microbiology 3, 64. doi: 10.3389/fmicb.2012.00064.
477
Wang, W., Sardans, J., Zeng, C., Zhong, C., Li, Y., Peñuelas, J., 2014. Responses of soil nutrient
478
concentrations and stoichiometry to different human land uses in a subtropical tidal wetland.
479
Geoderma 232–234(0), 459-470.
480
481
Wang, Z.P., DeLaune, R.D., Patrick, W.H., Masscheleyn, P.H., 1993. Soil Redox and pH Effects on
Methane Production in a Flooded Rice Soil. Soil Sci. Soc. Am. J. 57(2), 382-385.
482
WMO, 2014. Greenhouse gas bulletin, Geneva.
483
Ye, R., Jin, Q., Bohannan, B., Keller, J.K., McAllister, S.A., Bridgham, S.D., 2012. pH controls over
484
anaerobic carbon mineralization, the efficiency of methane production, and methanogenic
485
pathways in peatlands across an ombrotrophic–minerotrophic gradient. Soil Biology and
486
Biochemistry 54(0), 36-47.
487
Yin, S., Chen, D., Chen, L., Edis, R., 2002. Dissimilatory nitrate reduction to ammonium and
488
responsible microorganisms in two Chinese and Australian paddy soils. Soil Biology and
489
Biochemistry 34(8), 1131-1137.
30 | P a g e