Combined effects of carbon, nitrogen and phosphorus on CH4 production and denitrification in

wetland sediments

Sang Yoon Kim1, Annelies J. Veraart1, Marion Meima-Franke1 and Paul L.E. Bodelier1*

1Netherlands Institute of Ecology (NIOO-KNAW), department of microbial ecology, Wageningen, the

Netherlands

*Corresponding author: Paul L.E. Bodelier

Phone: +31 (0)317473485

Fax: +31 (0)317473675

E-mail address: p.bodelier@nioo.knaw.nl

Paper type: Regular article

1 | P a g e

1

2

3

4

5

6

7

8

9

10

11

12

13

1. Introduction

Methane (CH4) is the second most potent greenhouse gas in the atmosphere after carbon

dioxide (CO2) and has 34 times the global-warming potential of CO2 over a 100-year horizon (IPCC,

2013). Global atmospheric CH4 concentration has increased from pre-industrial level of 0.715 ppm to

1.824 ppm in 2013 (WMO, 2014). Wetlands, including rice paddies, are the largest natural source of

CH4 to the atmosphere, accounting for approximately 139 – 343 Tg CH4 yr-1. These ecosystems

contribute 32 – 47% to the total global CH4 emissions (Denman, 2007). CH4 is produced by a complex

microbial group which degrades organic matter by anaerobic methanogenesis (Conrad et al., 2007).

Methanogenesis is mediated by acetoclastic, hydrogenotrophic, and methylotrophic methanogens that

belong to the Euryarchaeota (Liu and Whitman, 2008).

Methanogenesis can be regulated by various factors including temperature (Conrad, 2002;

Glissman et al., 2004; Inglett and Inglett, 2013), pH (Wang et al., 1993; Ye et al., 2012), substrate

availability (O'Connor et al., 2010), and availability of electron acceptors (D'Angelo and Reddy, 1999).

Although a number of studies have been carried out to identify the main factors which control CH4

dynamics from wetlands, the effect of nutrients on CH4 dynamics is poorly understood. Many studies

point at nitrogen (N) as an important variable influencing CH4 cycles in wetland ecosystems (Bodelier,

2011). Effects of N addition can act directly on the methanogenic community, but may have indirect

effects, by stimulating the bacteria capable of denitrification. Denitrifiers and methanogens compete for

organic carbon (C), and furthermore denitrification produces toxic intermediates (NO, NO2, and N2O)

which can inhibit methanogenic archaea in wetland sediments, thereby reducing CH4 production (Roy

and Conrad, 1999). Although denitrification mitigates eutrophication effects in wetlands by reducing

availability of reactive nitrogen, incomplete denitrification also contributes to the emission of N 2O,

which is an even more potent greenhouse gas than CH4, with 300 times the global warming potential

of CO2 (IPCC, 2013).

Besides N, also phosphorus (P) can be an important regulating factor of methanogenesis. For

example, in Dutch drainage ditches, the water column PO43- concentration was found to be an

important predictor of CH4 emissions (Schrier-Uijl et al., 2011). Not only is P one of the most important

nutrients influencing the microbial activity, including decomposition processes (Cleveland et al., 2002),

but higher P concentrations also elevate microbial growth rates (Makino et al., 2003; Sterner and

2 | P a g e

14

15

16

17

18

19

20

21

22

23

24

25

26

27

28

29

30

31

32

33

34

35

36

37

38

39

40

41

42

Elser, 2002). Recently, Medvedeff et al. (2014) suggested that P addition stimulated methanogenic

activity in P-limited calcareous subtropical wetland soil, increasing the activity of methanogens directly

or indirectly via fermentative bacteria that produce methanogenic substrates.

Anthropogenic impacts on natural ecosystems have been increasing and largely influencing

the balance of essential nutrients such as C, N and P for decades (Howarth and Marino, 2006; Wang

et al., 2014). These changes in nutrient availability can impact the growth and activity of methanogens

and denitrifiers, and may significantly influence CH4 cycling. However, little is known about the

combined effects of C, N and P on CH4 cycling, and their impact on the interaction between

methanogenesis and denitrification remains unclear in wetland ecosystems.

The aim of this study was to investigate effects and possible interactions of N and P additions

on CH4 production and denitrification in wetland sediments. More specifically we tested the following

hypotheses: (i) sole P addition stimulates CH4 production due to elevated microbial growth rates

including methanogens in wetland sediments, (ii) combined N and P addition enhance denitrification

rates which lead to increased substrate competition between methanogens and denitrifiers, repressing

CH4 production from wetland sediments. To this end incubation studies were performed with sediment

derived from agricultural ditches in the Netherlands following methane production, denitrification,

functional gene abundance and sediment physico-chemistry. Because effects of N and P are expected

to act through competition for carbon, experiments were carried out with and without added C.

2. Materials and methods

2.1. Preparation and incubation of sediment slurries

The sediment samples were collected from the top layer (0-5 cm) of a drainage ditch sediment

in the spring of 2014 (Nigtevecht, The Netherlands; coordinates: 52° 16’ 41.92” N, 05° 01’ 40.11” E).

The drainage ditches are common water management practices in the Netherlands. These ditches

serve as important waterways and efficiently provide water to agricultural fields. The sediment pH was

neutral (6.92) and had comparatively high organic C content (63 g kg-1) as well as CH4 production

potential (106.9 nmol g-1 d.w hr-1) with low nutrient contents in pore water (Table 1). The sediment and

pore water characteristics are presented in Table 1. Prior to use, sediment samples were passed

3 | P a g e

43

44

45

46

47

48

49

50

51

52

53

54

55

56

57

58

59

60

61

62

63

64

65

66

67

68

69

70

through a stainless steel sieve (2 mm) by a wet-sieving method, and then stored at 4 °C until further

processing.

Anaerobic slurries were prepared (10 g fresh sediment and 13 ml amendment solution) in 150

ml serum bottles. The bottles were vigorously shaken by hand to homogenize the sediment slurries,

capped with sterilized butyl stoppers and then flushed with N2 for 1 hr. All bottles were incubated at

25 °C in the dark for 12 days on a shaker (120 rpm). In order to verify potential interacting effects of N,

P and C on CH4 production, we used a factorial design consisting of two C treatments (final conc. 0 or

10 mM C as CH3COOH), three P treatments (final conc. 0, 1, and 10 mM P as KH2PO4) and 3 N

treatments (final conc. 0, 2.5, and 5.0 mM N as KNO3), leading to a total of 18 combinations of C, N

and P. Each experiment was carried out in triplicate.The final concentrations of C, N and P, are

presented in Figure S1. N application level (0 - 5.0 mM, mean 2.5 mM) was chosen to resemble

agriculture fields that receive high N loads (< 10 mM as 200 kg N ha -1), as suggested by Roy and

Conrad (1999). In addition, P application level was determined by N to P ratios (0.25 - 5.0) based on

possible N and P input from fertilization by chemical fertilizer and manure applications which globally

ranged from 4.3 to 5.7 (Mean ca. 5) by estimating N and P balance for a century from 1950 to 2050

(Bouwman et al., 2013). C addition (10 mM) was chosen to achieve a ratio of 4 C: 1 N, which provides

sufficient C to the denitrifiers (Payne, 1981).

Note that although added nutrients were in the high range of drainage ditch water column

concentrations (Veraart, 2012), concentrations in the pore water, which are more relevant for this

study, frequently reach the mM range (Veraart et al., 2015), with our highest added concentrations

reflecting worst-case scenarios. Furthermore, nutrient concentrations are expected to increase with

fertilizer application and heavy agricultural runoff in wetland ecosystem after extreme weather events,

such as heavy rainfall, which due to global change is expected to occur more frequently in Western

Europe.

A control set of bottles (n = 3) was prepared to monitor the changes in headspace CH4

concentration during incubation and an additional control series (n = 3) was used for chemical analysis

of the pore water.

2.2. Measurements of CH4 production and chemical parameters

4 | P a g e

71

72

73

74

75

76

77

78

79

80

81

82

83

84

85

86

87

88

89

90

91

92

93

94

95

96

97

98

99

Methane concentrations were measured every day. Before sampling the headspace and pore

water, bottles were shaken vigorously to homogenize the sediment slurries. Gas samples (200 µl)

were taken using a gas-tight pressure-lock syringe flushed with N2. The CH4 in the headspace was

measured using an Ultra GC gas chromatograph (Interscience, the Netherlands) equipped with Rt-Q-

Bond (30 m, 0.32 mm, ID) capillary column and a flame ionization detector (FID). The temperature of

the column, injector and detector were adjusted to 80, 150, and 250 °C, respectively. Helium and H2

were used as the carrier and burning gases, respectively. Pore water samples (1 ml) were taken with

sterile disposable syringes equipped with long needles flushed with N2 to prevent O2 leakage that

might influence N mineralization processes such as ammonification and nitrification in the system. The

pore water was transferred to Eppendorf tubes (2 ml) and centrifuged for 15 min at 15,000 × g at 4 °C.

The supernatant was collected and stored at −20 °C until further analysis. Nutrient contents (NH4+,

NO3- and NO2

-) in the sediment were determined using an auto analyzer (QuAAtro, Seal analytical Inc.,

Beun de Ronde, Abcoude, The Netherlands). The pH in the slurries was measured directly after

addition of the amendment solutions and after incubation.

2.3. Denitrification potential

After the methanogenic incubations, denitrification potential measurements were conducted

using the acetylene inhibition method (Philippot et al., 2013; Qin et al., 2012). Briefly, 5g of fresh slurry

from the bottles after incubation was transferred to new bottles (150 ml) and 7.5 ml of a solution

containing KNO3 (1 mM) and glucose (1 mM) was added. The bottles were sealed and capped with

sterilized butyl stoppers and then flushed with N2 for 10 min. Acetylene was added (10% v/v of the

headspace) and the bottles were incubated at 25 °C on a shaker (120 rpm). Nitrous oxide (N2O) in the

headspace was measured after 24 hrs. of incubation using a TRACE 1300 (Thermo Scientific, the

Netherlands) equipped with HS-Q (1m, 2.0 mm, ID) packed column fitted with an electron capture

detector (ECD). The temperature of the column and injector was 90 °C and the detector was set at

350 °C. N2 gas was used as the carrier.

2.4. Quantification of mcrA and nirS gene copy numbers in sediment slurries

5 | P a g e

100

101

102

103

104

105

106

107

108

109

110

111

112

113

114

115

116

117

118

119

120

121

122

123

124

125

126

127

After incubation, DNA was extracted from 0.10 g of freeze-dried sediment following a slightly

modified protocol based on the FastDNA spin kit for soil (MP Biomedicals, Solon, OH) previously

described in detail in Wang et al. (2012). Nucleic acids were routinely quantified using a NanoDrop

1000 spectrophotometer (Thermo) before quantitative PCR (qPCR). The copy numbers of the mcrA

gene, encoding the methyl coenzyme-M reductase were used as proxy for methanogenic abundance

and the copy numbers of the nirS gene, encoding the cytochrome cd1 nitrite reductase were used as

proxy for denitrifier abundance. Primer sets of mlas/mcrA-rev were used for mcrA (Steinberg and

Regan, 2008) and nirScd3af/nirSR3cd for nirS (Throbäck et al., 2004). Real-time qPCR was

performed in a Rotor-Gene Q real-time PCR cycler (Qiagen, the Netherlands). Briefly, qPCR reaction

(total volume 20 µl) for mcrA gene contained 10 µl of reaction mixtures 2X SensiFAST SYBR

(BIOLINE, the Netherlands), 3.5 µl of forward primers (4 pmol µl-1), 3.5 µl reverse primers (5 pmol µl-1),

1 µl Bovine Serum Albumin (5 mg ml-1; Invitrogen, the Netherlands), and 2 µl diluted template DNA (2

ng µl-1). Amplification was carried out as follows: for the mcrA gene, initial denaturation at 95 °C for 3

min, followed by 45 cycles of denaturation at 95 °C for 10 sec, annealing at 60 C for 10 sec, and

extension at 72 °C for 25 sec. qPCR reaction (total volume 20 µl) for nirS gene contained 10 µl of

reaction mixtures 2X SensiFAST SYBR (BIOLINE, the Netherlands), 2 µl of forward primers (5 pmol µl -

1), 2 µl reverse primers (5 pmol µl-1), 1 µl Bovine Serum Albumin (5 mg ml-1; Invitrogen, the

Netherlands), and 2.5 µl diluted template DNA (2 ng µl -1). Amplification was carried out as follows: for

the nirS gene, initial denaturation at 95 °C for 3 min, followed by 40 cycles of denaturation at 95 °C for

10 sec, annealing at 60 °C for 10 sec, and extension at 72 °C for 20 sec and 86 °C for 5 sec. In each

qPCR, the fluorescence signal was obtained at 72 °C after each cycle, and melt curves were obtained

from 70 °C to 99 °C (1 °C temperature by rising). Amplicon specificity was determined from the melt

curve. To avoid the inhibitory effects of substances co-extracted with the DNA, amplification of serial

dilutions was carried out for the slurry samples in each treatment.

2.5. Statistical analysis

Statistical analyses were conducted using R studio software (ver.2.6.0). Determination of

differences between parameters was performed through three-way analysis of variance (ANOVA)

including N, P and C additions and their interactions.

6 | P a g e

128

129

130

131

132

133

134

135

136

137

138

139

140

141

142

143

144

145

146

147

148

149

150

151

152

153

154

155

156

3. Results

3.1. Effects of C, N and P on CH4 production and potential denitrification

3.1.1 CH4 production and its lag phase

C, N, P and their interaction all significantly affected CH4 production and its lag phase (Figures

1 and 2, Table 2, three-way ANOVA, see Table 3). Overall, we observed most CH4 in the headspace

of samples where no N was added, of which samples with added C had the highest final CH4

concentrations (113.2 µM CH4). When C was added, effects of N and P clearly interacted: the lowest

CH4 production (1.14 µM CH4) was found in the samples where the highest P (10 mM) and N (5 mM)

were added. Combined C and P addition at 10 mM increased the lag phase of CH4 production,

especially in combination with N addition (Figure 1 and Table 2).

3.1.2. Denitrification potential

Denitrification potential was on average 28.2 nmol N g -1 d.w-1, and showed interacting effects

of C, N and P (Figure 2, three-way ANOVA, see Table 3). When no P was added, addition of N and C

moderately stimulated denitrification. With 1 mM P added, C addition moderately enhanced

denitrification, but for the treatment without C the effect of N is not clear. Interestingly, a very different

pattern can be seen for the highest P addition (10 mM), where there seems to be an additive inhibitory

effect of N and C addition: the lowest denitrification potential (8.28 nmol N g-1 d.w-1) was found when P,

N and C were added.

3.2. Changes in biogeochemical properties

3.2.1. Changes of NO3-, NO2

-, and NH4+ in pore water

NO3- was consumed within 48 hr in sediment slurries without P addition, but decreased slower

in slurries where P was added, irrespective of C addition. In sediment slurries not amended with N and

C, no NO2- was detected during the incubation period. NO2

- was only observed in the sediment to

which both C and N were added, and was higher in the slurries with added P (10 mM). For example,

there was 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, and 10 mM). The NO2- peak occurs when NO3

- levels are still decreasing, and vanishes with

disappearing NO3-. Interestingly, although without P addition, only small fluctuations in NH4

+ occurred

7 | P a g e

157

158

159

160

161

162

163

164

165

166

167

168

169

170

171

172

173

174

175

176

177

178

179

180

181

182

183

184

185

186

in all treatments, when C, N and P were added in combination, an increase in NH 4+

above the baseline

variation can be observed (especially at N 2.5 mM), which later decreases again. Overall, NH4+

concentrations were maintained at a higher level in P-added slurries (0.036 - 0.103 mM on average)

than in slurries where no P was added (0.020 - 0.034 mM on average) (Figure 3).

3.2.2. Changes of pH

C addition decreased the pH of the sediment slurries, but during incubation pH in these

slurries neutralized again (Figure S2). P addition also acidified the slurries, with lower pH in P10

slurries (ca. 6.30) than in P0 and P1 slurries (ca. 6.73), and these effects remained after incubation.

3.2.3. mcrA and nirS gene copy numbers

The abundances of the mcrA gene and nirS gene were determined by qPCR after slurry

incubation (Figure 4). The mcrA gene ranged from 6.78 × 107 to 1.10 × 108 and nirS gene copy

numbers ranged from 5.21 × 106 to 6.62 × 107 copies g-1 dry sediment. Unexpectedly, the copy

numbers of mcrA genes were higher in samples without C addition but increased with P addition

except N0 treatment. The effect of N addition on mcrA gene copy was not observed under C addition.

In particular, N addition gradually increased nirS gene copy numbers in slurries with added P or C, but

copy numbers decreased when P and C were both added. We found a weak correlation but significant

negative exponential relations between CH4 production and mcrA gene copy numbers (R2=0.164,

P=0.051). A positive exponential correlation between nirS gene copy numbers and denitrification

potential (R2=0.288, P<0.01) was also observed.

3.3. Relationship between CH4 production and denitrification potential

P additions (P1 and P10) without added N showed similar patterns as the control without

added P, and did not influence CH4 production and denitrification potential, irrespective of C addition

(Figure 5). N additions without added P decreased CH4 production but increased denitrification

potential in sediments with and without added C. Combined N and P addition showed additive effects

on CH4 production and denitrification potential. However, the processes had a different response to C

addition: when C was added, CH4 production and denitrification were more strongly inhibited at high

N:P ratios (2.5 - 5.0) than at low N:P ratios (0.25 - 0.5).

8 | P a g e

187

188

189

190

191

192

193

194

195

196

197

198

199

200

201

202

203

204

205

206

207

208

209

210

211

212

213

214

215

216

4. Discussion

We hypothesized that sole and combined effects of N and P on methanogenesis and

denitrification in wetlands sediments act through competition for C. Experiments were carried out with

and without added C to measure CH4 production potential and potential denitrification as well as

biogeochemical properties during slurry incubations. The present study shows combined effects of N

and P on CH4 and N cycles in wetland sediments, and indicates that these effects may indeed occur

through competition for C - albeit in a more complex interaction than initially hypothesized - which will

be discussed in the following sections.

4.1. Effects of C, N and P on CH4 production and denitrification

Our study showed that CH4 production was significantly affected by C addition in sediment

slurries. It is well-known that acetate can be directly utilized by acetogenic methanogens to produce

CH4 under anaerobic conditions (Conrad, 2007). This stimulation was not reflected in the mcrA gene

abundance, for which we have no definite explanation other than referring to other studies which also

found no relation between gene abundance and methane production (Cadillo‐Quiroz et al., 2006;

Galand et al., 2003). Possibly, acetate and phosphate have formed acetylphosphate, which may have

affected enzyme functioning in the metabolic pathways and qPCR process.

N additions inhibited CH4 production (Figure 2). Its effect increased with increasing N

concentration. The inhibitory effect of N compounds on the methanogenic microbial community can be

caused by denitrification intermediates (NO, NO2-) that are toxic for methanogenic archaea, or by

competition between denitrifiers and methanogens for acetate (Kluber and Conrad, 1998; Roy and

Conrad, 1999). NO2- was only produced in excess when C and N were added, and this effect was

much stronger under high P (Figure 3). This result indicates that the increased C/N ratio influenced the

first denitrification steps and that the N:P ratio can also affect NO 2- accumulation. The accumulation of

NO2- stops when NO3

- levels decrease, which is also expected from theoretical denitrification

energetics (van de Leemput et al., 2011).

9 | P a g e

217

218

219

220

221

222

223

224

225

226

227

228

229

230

231

232

233

234

235

236

237

238

239

240

241

242

243

Roy and Conrad (1999) speculated that competition for substrate between denitrifiers and

methanogens is not the main mechanism of suppression of methanogenesis in rice soil. However, we

observed inhibitory effects of N addition on CH4 production and lag phase in slurries where no C was

added (Figure 2). In these slurries, no NO2- accumulation was observed (Figure 3), which indicates

that the inhibitory effect is mainly due to the competition for C between methanogens and other

microorganisms.

Our study showed that CH4 production was not directly influenced by P addition in sediment

slurries, irrespective of C addition. However, P addition slowed down NO3- uptake by denitrifiers, which

can increase C availability for methanogens by decreasing competition. Therefore, we expected CH4

production to be increased during the delayed NO3- uptake. Different to our expectation, CH4

production was significantly lower when N and P were added, irrespective of C addition (Figure 2). It

appears that low N:P ratios decreased denitrification. A negative correlation between P and

denitrification was also observed in the sediment of C-poor shallow lakes (Veraart, 2012). On the other

hand, seasonal positive correlations between total P and denitrification were observed in a meta-

analysis of denitrification across aquatic ecosystems (Piña-Ochoa and Álvarez-Cobelas, 2006).

However, P-release from the sediment and denitrification occur under similar anoxic conditions,

confounding ecosystem observations, and making negative correlations between P and denitrification

even more interesting. The potential inhibitory effect of P on denitrification may be direct or indirect.

Potential direct effects are not well understood, but may be through P-sensitivity of the denitrifiers due

to adaptation to the low prevailing P conditions in this sediment. Indirect effects through enhanced

competition for C and N with other bacterial populations are also possible. We used copy numbers of

nirS, reflecting the denitrifier community containing the cytochrome cd1 nitrite reductase, as a proxy for

denitrifier abundance. This may give an incomplete picture, because part of the denitrifiers will use

nirK rather than nirS, and mere gene presence does not directly reflect functional activity. Nonetheless

abundance of nirS positively correlated with potential denitrification (Figure S3). Copy numbers of nirS

were also highest when both P and N were present, which may hint at an N and P co-limitation of the

denitrifying community. However, much like mcrA copy numbers, this effect of N and P on nirS copy

numbers was only observed when no C was added.

10 | P a g e

244

245

246

247

248

249

250

251

252

253

254

255

256

257

258

259

260

261

262

263

264

265

266

267

268

269

270

271

272

4.2. Competition between denitrification, DNRA and methanogenesis

C addition may have shifted the competitive balance of different nitrate respiring and C

degrading communities. Possibly, communities carrying out DNRA (dissimilatory nitrate reduction to

ammonium) may have been better competitors for acetate under N and P addition, suppressing the

abundance of nirS-denitrifiers as well as methanogenic activity. DNRA-communities compete with

denitrifiers for C and N, as these pathways occur under similar conditions (Francis et al., 2007; Scott

et al., 2008; Tiedje, 1988; Yin et al., 2002). The C:N ratio is a key determinant of the competitive

outcome between both processes, favoring DNRA at high C availability (Smith, 1982), in line with our

findings of lower denitrification under combined N, P and C addition. Our ‘DNRA-competition’-

hypothesis is further supported by the observed increase in NH4+ concentrations in slurries with

combined N, P and application (Figure 2). However, we estimate that about 10% of the NO3- was

converted to NH4+ - rather than gaseous nitrogen - in these slurries, leaving part of the difference in

denitrification potential unexplained. A closed mass balance or isotope tracer approach would be

optimal to accurately trace N-conversions.

Interestingly, combined C and P addition seems to enhance initial ammonification and DNRA,

suggesting that these steps of the N cycle are P limited in this sediment. Clearly, further studies are

needed to test P-limitation of ammonification, DNRA, and denitrification and consequently its

implications for CH4 cycling. Isotope tracing methods based on single cells (Krause et al., 2014) such

as stable isotope tracers and nano-SIMS, in combination with bacterial community analyses to identify

shifts between dominant populations and N-respiring pathways (Kraft et al., 2014) will be promising

ways to study this.

5. Conclusions

We have summarized the main findings of this study in a conceptual scheme depicted in

Figure 6. The effects of N, P and C additions and the interaction between methanogenesis and

denitrification turned out to be more complicated than hypothesized. Instead of simply stimulating

growth of methanogens or denitrifiers by N or P additions we observed an interactive effect of C, N

and P which may be explained by a modulation of electron flow towards DNRA, consuming electron

11 | P a g e

273

274

275

276

277

278

279

280

281

282

283

284

285

286

287

288

289

290

291

292

293

294

295

296

297

298

299

300

donors in the process. Despite the fact that this hypothesis still has to be verified experimentally it is

safe to conclude that P might play an important modulating role in carbon degradation and C-N cycle

interactions in wetland sediments with possible consequences for greenhouse gas emissions from

these ecosystems.

Acknowledgements

This work was supported a grant (823.001.008) of the Netherlands Organisation for Scientific

Research (NWO). This work was supported by the National Research Foundation of Korea (NRF)

grant funded by the Korea government (MSIP) (No. NRF-2013R1A2A2A07068946). We thank Anne

Steenbergh for help with the statistical analysis. This publication is publication no. xxxx of the

Netherlands Institute of Ecology.

12 | P a g e

301

302

303

304

305

306

307

308

309

310

311

Table 1. Characteristics of the sediment and pore water before the experiment.

Parameters Value

Sediment

pH

Total C (g kg-1)

Total N (g kg-1)

C:N ratio

Potential CH4 production (nmol g-1 d.w hr-1)

Pore water

Electrical conductivity (µS cm-1)

Dissolved organic C (mg L-1)

Dissolved inorganic C (mg L-1)

NH4+ (mg L-1)

PO43- (mg L-1)

N:P ratio

6.92

63.0

5.42

11.6

106.9

853

14.0

3.8

0.43

2.0

0.51

13 | P a g e

312

313

Table 2. Variables of methanogenic potentials in sediment slurries of different nutrient addition

treatments.

Treatment Variables

Lag phase (hr)* CH4 production rate (umol g-1 hr-1) R2***

Without

C addition

P0 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

P1 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

P10 N0 12.7 ± 0.90 0.038 ± 0.002 0.998

N2.5 - BD** -

N5.0 - BD -

With C addition P0 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

P1 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

P10 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 -

* Initiation of linear phase of CH4 production in the data using four time points (extrapolated from linear

regressions).

** BD means below detection limits (CH4 production rate < 0.001).

*** R2 values were estimated from CH4 production in the data.

14 | P a g e

314

315

316

317

318

319

Table 3. Three-way ANOVA results. All variables except pH were log-transformed to improve normality.

variable Total CH4

production

CH4 lag-phase Potential

denitrification rate

CH4 production

rate

mcrA nirS pH

(before incubation)

pH

(after incubation)

F P F P F P F P F P F P F P F P

C x N x P 19.91 <0.001 6.392 0.016108.3

3<0.001 0.646 0.427 6.341 0.018 25.164 <0.001 0.119 0.7317 6.69 0.013

N x P 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

C x P 15.60 <0.001 42.418 <0.001327.3

1<0.001 20.37 <0.001 0.501 0.485 52.418 <0.001 2.860 0.098 0.024 0.877

C x N 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 <0.001C 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 0.00438.49

5<0.001 93.71 <0.001

P 63.22 <0.001 1.715 0.199170.6

1<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

321

Figure legends

Figure 1. Changes in CH4 concentrations in different nutrient-amended slurries during incubation with

C or without C added. Bars represent standard deviations (n=3). Each panel shows a different P

addition level and the lines per panel show different N addition levels.

Figure 2. Total CH4 production and denitrification potential after incubation. Bars represent standard

deviations (n=3).

Figure 3. Changes in concentrations of nitrate (NO3-), nitrite (NO2

-), and ammonium (NH4+) in pore

water in different nutrient-amended slurries during incubation with C or without C addition. Bars

represent standard deviations (n=3).

Figure 4. Abundances of the mcrA and nirS genes in different nutrient-amended sediments after

incubation. Bars represent standard deviations (n=3).

Figure 5. Effect of N:P ratio on CH4 production and denitrification.

Figure 6. Conceptual scheme of observed and hypothetical effects of the applied treatments in this

study.

16 | P a g e

322

323

324

325

326

327

328

329

330

331

332

333

334

335

Figure 1.

17 | P a g e

336

337

338

339

Figure 2.

18 | P a g e

340

341

342

343

Figure 3.

19 | P a g e

344

345

346

Figure 4.

20 | P a g e

347

348

349

350

Figure 5.

21 | P a g e

351

352

353

354

Figure 6.

22 | P a g e

355

356

357

358

Supporting Information Legends

Figure S1. Experimental scheme for nutrient and substrate additions.

Figure S2. Initial and final pH of incubation in different nutrient-amended sediments. Bars represent

standard deviations (n=3).

Figure S3. Relationships between functional gene abundances and their activities.

23 | P a g e

359

360

361

362

363

Figure S1.

24 | P a g e

• N- (0 mM) • N+ (2.5 mM)• N++ (5.0 mM)

• N- (0 mM) • N+ (2.5 mM)• N++ (5.0 mM)

P++(10 mM)

P+ (1 mM)

• N- (0 mM) • N+ (2.5 mM)• N++ (5.0 mM)

P-(0 mM)

• N- (0 mM) • N+ (2.5 mM)• N++ (5.0 mM)

• N- (0 mM) • N+ (2.5 mM)• N++ (5.0 mM)

Wetlandsediment

C +(10 mM)

C -(0 mM )

P++(10 mM)

P+(1 mM)

• N- (0 mM) • N+ (2.5 mM)• N++ (5.0 mM)

P-(0 mM)

364

365

366

367

Figure S2.

25 | P a g e

368

369

370

371

Figure S3.

26 | P a g e

372

373

374

375

REFERENCES

Bodelier, P.L., 2011. Interactions between nitrogenous fertilizers and methane cycling in wetland and

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.,

Rufino, M.C., Stehfest, E., 2013. Exploring global changes in nitrogen and phosphorus cycles

in agriculture induced by livestock production over the 1900–2050 period. Proceedings of the

National Academy of Sciences 110(52), 20882-20887.

Cadillo‐Quiroz, H., Bräuer, S., Yashiro, E., Sun, C., Yavitt, J., Zinder, S., 2006. Vertical profiles of

methanogenesis and methanogens in two contrasting acidic peatlands in central New York

State, USA. Environ Microbiol 8(8), 1428-1440.

Cleveland, C.C., Townsend, A.R., Schmidt, S.K., 2002. Phosphorus Limitation of Microbial Processes

in Moist Tropical Forests: Evidence from Short-term Laboratory Incubations and Field Studies.

Ecosystems 5(7), 0680-0691.

Conrad, R., 2002. Control of microbial methane production in wetland rice fields. Nutr Cycl Agroecosys

64(1-2), 59-69.

Conrad, R., 2007. Microbial ecology of methanogens and methanotrophs. Adv Agron 96, 1-63.

Conrad, R., Chan, O., Claus, P., Casper, P., 2007. Characterization of methanogenic Archaea and

stable isotope fractionation during methane production in the profundal sediment of an

oligotrophic lake (Lake Stechlin, Germany). Limnol Oceanogr 52(4), 1393.

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,

D. Hauglustaine, C. Heinze, E. Holland, D. Jacob, U. Lohmann, S Ramachandran, P.L. da

Silva Dias, S.C. Wofsy and X. Zhang, 2007: Couplings Between Changes in the Climate

System and Biogeochemistry. In: Climate Change 2007: The Physical Science Basis.

Contribution of Working Group I to the Fourth Assessment Report of the Intergovernmental

Panel on Climate Change [Solomon, S., D. Qin, M. Manning, Z. Chen, M. Marquis, K.B.

Averyt, M.Tignor and H.L. Miller (eds.)]. Cambridge University Press, Cambridge, United

Kingdom and New York, NY, USA.

27 | P a g e

376

377

378

379

380

381

382

383

384

385

386

387

388

389

390

391

392

393

394

395

396

397

398

399

400

401

402

403

404

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

community structure in the sediment of eutrophic Lake Dagow: effect of temperature. Microb

Ecol 48(3), 389-399.

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

the Fifth Assessment Report of the Intergovernmental Panel on Climate Change [Stocker,

T.F., D. Qin, G.-K. Plattner, M. Tignor, S.K. Allen, J. Boschung, A. Nauels, Y. Xia, V. Bex and

P.M. Midgley (eds.)]. Cambridge University Press, Cambridge, United Kingdom and New

York, NY, USA, 1535 pp.

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.

Kraft, B., Tegetmeyer, H.E., Sharma, R., Klotz, M.G., Ferdelman, T.G., Hettich, R.L., Geelhoed, J.S.,

Strous, M., 2014. The environmental controls that govern the end product of bacterial nitrate

respiration. Science 345(6197), 676-679.

Krause, S., Le Roux, X., Niklaus, P.A., Van Bodegom, P.M., Lennon, J.T., Bertilsson, S., Grossart, H.-

P., Philippot, L., Bodelier, P.L., 2014. Trait-based approaches for understanding microbial

biodiversity and ecosystem functioning. Frontiers in Microbiology 5, 251

210.3389/fmicb.2014.00251

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?

Growth rate and resource dependence of bacterial C : N : P stoichiometry. Funct Ecol 17(1),

121-130.

28 | P a g e

405

406

407

408

409

410

411

412

413

414

415

416

417

418

419

420

421

422

423

424

425

426

427

428

429

430

431

432

433

434

Medvedeff, C.A., Inglett, K.S., Inglett, P.W., 2014. Evaluation of direct and indirect phosphorus

limitation of methanogenic pathways in a calcareous subtropical wetland soil. Soil Biology and

Biochemistry 69(0), 343-345.

O'Connor, F.M., Boucher, O., Gedney, N., Jones, C.D., Folberth, G.A., Coppell, R., Friedlingstein, P.,

Collins, W.J., Chappellaz, J., Ridley, J., Johnson, C.E., 2010. Possible role of wetlands,

permafrost, and methane hydrates in the methane cycle under future climate change: A

review. Reviews of Geophysics 48(4), RG4005.

Payne, W.J., 1981. Denitrification. John Wiley & Sons Inc.

Philippot, L., Spor, A., Henault, C., Bru, D., Bizouard, F., Jones, C.M., Sarr, A., Maron, P.A., 2013.

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

suppression of methane production by nitrate in anoxic rice field soil. Fems Microbiol Ecol

28(1), 49-61.

Schrier-Uijl, A., Veraart, A., Leffelaar, P., Berendse, F., Veenendaal, E., 2011. Release of CO2 and

CH4 from lakes and drainage ditches in temperate wetlands. Biogeochemistry 102(1-3), 265-

279.

Scott, J.T., McCarthy, M., Gardner, W., Doyle, R., 2008. Denitrification, dissimilatory nitrate reduction

to ammonium, and nitrogen fixation along a nitrate concentration gradient in a created

freshwater wetland. Biogeochemistry 87(1), 99-111.

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

an acidic, oligotrophic fen and an anaerobic digester treating municipal wastewater sludge.

Applied and environmental microbiology 74(21), 6663-6671.

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

435

436

437

438

439

440

441

442

443

444

445

446

447

448

449

450

451

452

453

454

455

456

457

458

459

460

461

462

463

464

Throbäck, I.N., Enwall, K., Jarvis, Å., Hallin, S., 2004. Reassessing PCR primers targeting nirS, nirK

and nosZ genes for community surveys of denitrifying bacteria with DGGE. Fems Microbiol

Ecol 49(3), 401-417.

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. Predicting

microbial nitrogen pathways from basic principles. Environ Microbiol 13(6), 1477-1487.

Veraart, A.J., 2012. Denitrification in Ditches, Streams and Shallow Lakes. Wageningen University.

Veraart, A.J., Steenbergh, A., Ho, A., Kim, S.Y., Bodelier, L.E.P., 2015. Beyond nitrogen: The

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.

Spatial patterns of iron- and methane-oxidizing bacterial communities in an irregularly flooded,

riparian wetland. Frontiers in Microbiology 3, 64. doi: 10.3389/fmicb.2012.00064.

Wang, W., Sardans, J., Zeng, C., Zhong, C., Li, Y., Peñuelas, J., 2014. Responses of soil nutrient

concentrations and stoichiometry to different human land uses in a subtropical tidal wetland.

Geoderma 232–234(0), 459-470.

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.

WMO, 2014. Greenhouse gas bulletin, Geneva.

Ye, R., Jin, Q., Bohannan, B., Keller, J.K., McAllister, S.A., Bridgham, S.D., 2012. pH controls over

anaerobic carbon mineralization, the efficiency of methane production, and methanogenic

pathways in peatlands across an ombrotrophic–minerotrophic gradient. Soil Biology and

Biochemistry 54(0), 36-47.

Yin, S., Chen, D., Chen, L., Edis, R., 2002. Dissimilatory nitrate reduction to ammonium and

responsible microorganisms in two Chinese and Australian paddy soils. Soil Biology and

Biochemistry 34(8), 1131-1137.

30 | P a g e

465

466

467

468

469

470

471

472

473

474

475

476

477

478

479

480

481

482

483

484

485

486

487

488

489

490