Science of the Total Environment xxx (2017) xxx–xxx

STOTEN-24399; No of Pages 10

Contents lists available at ScienceDirect

Science of the Total Environment

j ourna l homepage: www.e lsev ie r .com/ locate /sc i totenv

Seasonal variations of nitrous oxide fluxes and soil denitrification rates insubtropical freshwater and brackish tidal marshes of the MinRiver estuary

Xuming Wang a,b,c,1, Minjie Hu a,1, Hongchang Ren a, Jiabing Li d, Chuan Tong a,e,⁎, Ronald S. Musenze f,g,⁎a School of Geographical Sciences, Fujian Normal University, Fuzhou, Chinab State Key Laboratory of Urban and Regional Ecology, Research Center for Eco-Environmental Sciences, Chinese Academy of Sciences, Beijing, Chinac University of Chinese Academy of Sciences, Beijing, Chinad School of Environmental Science and Engineering, Fujian Normal University, Fuzhou, Chinae Key Laboratory of Humid Sub-tropical Eco-geographical Process of Ministry of Education, Fujian Normal University, Fuzhou, Chinaf Air Quality Monitoring Unit, Department of Science, Information Technology and Innovation, Queensland Government, Brisbane, Australiag Department of Civil and Environmental Engineering, College of Engineering, Design, Art and Technology, School of Engineering, Makerere University, Kampala, Uganda

H I G H L I G H T S G R A P H I C A L A B S T R A C T

• N2O fluxes were significantly higher infreshwater marshes than in brackishmarshes.

• Soil denitrification rates were higher infreshwater than in brackish-watermarshes.

• Both N2O flux and denitrification rateshowed strong seasonal variability.

• N2O fluxes had strong negative correla-tions with salinity in tidal marshes.

• Temperature, salinity and ammoniaconcentration were key controls ofdenitrification.

⁎ Corresponding authors.E-mail addresses: tongch@fjnu.edu.cn (C. Tong), r.mus

1 Co first authors who contributed equally to this work

https://doi.org/10.1016/j.scitotenv.2017.10.1750048-9697/© 2017 Elsevier B.V. All rights reserved.

Please cite this article as: Wang, X., et al., Sebrackish tidal marshes of..., Sci Total Environ

a b s t r a c t

a r t i c l e i n f o

Article history:Received 1 June 2017Received in revised form 15 October 2017Accepted 17 October 2017Available online xxxx

Editor: D. Barcelo

Estuarine tidal marshes provide favorable conditions for nitrous oxide (N2O) production. Saltwater intrusioncaused by sea-level rise would exert complex effects on the production and emission of N2O in estuarine tidalmarshes; however, few studies have been conducted on its effects onN2O emissions. Salinity gradients are a com-mon occurrence in estuarine tidalmarshes. Studies on production and emission of N2O in tidal marshes with dif-ferent salinities may elucidate the impact of saltwater intrusion on the emission of greenhouse gases. This studyexplores the seasonal variations of N2O fluxes and soil denitrification rates in freshwater (Daoqingzhouwetland)and brackish (Shanyutan wetland) tidal marshes dominated by Cyperus malaccensis var. brevifolius (shichitomatgrass) in theMin River estuary, southeastern China. N2O fluxes in bothmarshes showed strong temporal var-iability. The highest N2O fluxes were observed in the hot and wet summer months, whereas the lowest fluxeswere observed in the cold winter and autumn months. N2O fluxes from the freshwater marsh (48.81 ± 9.01μg m−2 h−1) were significantly higher (p b 0.05) than those from the brackish-water marsh (27.69 ± 4.01μg m−2 h−1). Soil denitrification rates showed a similar temporal pattern, with the highest rates observed insummer and the lowest in winter. Similarly, soil denitrification rates were significantly higher (p b 0.05) in thefreshwatermarsh (32.72± 19.15 μmol Nm−2 h−1) than in the brackish-watermarsh (4.97± 2.64 μmol Nm−2-

h−1). Temperature and the salinity, sulfate (SO42−), and ammonia nitrogen (NH4

+-N) concentrations of the

Keywords:Nitrous oxide emissionDenitrificationSalinity gradientBrackish tidal marshSaltwater intrusionSubtropical estuary

enze@cedat.mak.ac.ug (R.S. Musenze)..

asonal variations of nitrous oxide fluxes and soil denitrification rates in subtropical freshwater and(2017), https://doi.org/10.1016/j.scitotenv.2017.10.175

2 X. Wang et al. / Science of the Total Environment xxx (2017) xxx–xxx

Please cite this article as: Wang, X., et al., Sebrackish tidal marshes of..., Sci Total Environ

overlying water were key factors affecting soil denitrification rates. N2O fluxes and soil denitrification rates dem-onstrated negative correlations with salinity and SO4

2− concentrations in both marshes. The results indicate thatincreased seawater intrusion would reduce N2O emissions from estuarine tidal wetlands and exert a negativefeedback on the climate system.

© 2017 Elsevier B.V. All rights reserved.

1. Introduction

Nitrous oxide (N2O) is a potent greenhouse gas with a globalwarming potential (GWP) 298 times that of carbon dioxide (CO2) overa 100-year time horizon (Denman et al., 2007) and a long atmosphericlifetime (Bauza et al., 2002). Although atmospheric concentrations ofN2O are only about 10% of CO2 concentrations, the damage caused tothe global environment by N2O emissions is very significant. N2O isalso currently recognized as the single most important ozone-depletingsubstance (Ravishankara et al., 2009).

Sea-level rise caused by climate change is of great scientific interest(Wigley and Raper, 1992; Rahmstorf, 2007; Carson et al., 2016). TheFifth Assessment Report (AR5) of the Intergovernmental Panel on Cli-mate Change (IPCC) reported a global average sea level rise of 0.19 mbetween 1901 and 2010, with saltwater intrusion occurring widely incoastal areas (Myhre et al., 2013). Saltwater intrusion may pose signifi-cant ecological threats to fresh- and brackish-water marsh ecosystems(Knowles, 2002; Herbert et al., 2015).

Many tidal marshes with various salinity levels develop alongestuaries at different tidal extents, influenced by both freshwater riverflows and ocean tides. Estuarine tidal marshes, which are gateways ofterrestrial nitrogen discharge into the ocean, provide favorableconditions (e.g., alternating dry and wet environments/sediments) forthe production and emission of N2O (Galloway et al., 2003;Moseman-Valtierra et al., 2011; Wang et al., 2016). Sediment denitrifi-cation,which produces N2O as an intermediate product, is amajor path-way of nitrogen removal in tidal marshes (Jordan et al., 2011;Morley etal., 2014). Salinity in tidalmarshes is an important environmental factorinfluencing the generation and emissions of N2O via the sediment deni-trification pathway. Salinity affects (1) the enhancement of ammoniumrelease from sediments through physical and chemical processes; (2)the nitrogen release from sediment through physiological processes,and (3) microbial activity (by both nitrifiers and denitrifiers) changingthe rate of dissimilatory nitrate reduction to ammonium (Giblin et al.,2010; Wang et al., 2014). However, the effects of salinity on N2O emis-sions and sediment denitrification are complex.

Studies evaluating the effect of salinity changes on N2O emissions innatural environments remain limited, and conflicting data have beenreported. Some studies have reported the inhibition of denitrificationand a consequent reduction in N2O emissions (Seo et al., 2008;Santoro, 2010; Chauhan et al., 2015; Osborne et al., 2015); however,the conclusions of those studies were inconsistent (Fear et al., 2005;Marton et al., 2012). Previously, little attention was paid to changes insediment nitrogen transformations and N2O emissions in environmentsexperiencing small salinity changes, such as tidal marshes (Franklin etal., 2016). However, investigation of these changes would add valuableknowledge about climate feedback associated with sea-level rise in thecoming decades (Woodroffe and Murray-Wallace, 2012).

Simultaneous field measurements of N2O fluxes in both freshwaterand brackish-water tidal marshes would be instrumental for enrichingthe body of knowledge about greenhouse gas emissions in subtropicalestuarine systems where data are still scarce (Musenze et al., 2014,2015). This study presents a quantitative spatial and temporal eval-uation of soil denitrification rates and N2O fluxes in both freshwaterand brackish-water tidal marshes dominated by Cyperus malaccensisvar. brevifolius (shichito matgrass) in the Min River estuary, southeast-ern China. Shichito matgrass is a common coastal wetlands species,

asonal variations of nitrous ox(2017), https://doi.org/10.10

especially in Fujian and Guangdong provinces. The objectives of thisstudy are to: (1) determine the effects of small increases in salinity onN2O fluxes and denitrification rates in subtropical estuarine tidalmarsh ecosystems and (2) establish key environmental controls forN2O emissions in estuarine marshes with narrow salinity gradients inthe subtropics.

2. Materials and methods

2.1. Site description

This study was conducted in the Min River estuary with measure-ment sites located in the Daoqingzhouwetland (a freshwater tidal wet-land) and the Shanyutan wetland (a brackish-water tidal wetland). Thefreshwater shichito matgrass marsh site was located in the southwest-ern part of the Daoqingzhou wetland (25°57′21″N, 119°24′25″E), andthe brackish-water shichitomatgrassmarsh sitewas located in themid-dle-western part of the Shanyutan wetland (26°01′46″N, 119°37′31″E)(Fig. 1). The climate of this area is subtropical with hot, humid summersand dry, coldwinters. Themean annual temperature of the estuarine re-gion is 19.85 °C, and themean annual precipitation is 1905mm. Shichitomatgrass is a typical native species in both the Daoqingzhou andShanyutan wetlands. In the Min River estuary, the tides are consideredtypical semidiurnal tides; the soil surface is generally submerged for ap-proximately 7 h over a 24 h cycle (Tong et al., 2013). The average salin-ity of the tidal water (measured from November 2013 to October 2014)was 0.20 ± 0.02‰ and 3.79 ± 1.35‰ in the Daoqingzhou andShanyutan wetlands, respectively (see Table 2 for variation in seasonalvalues). There were no significant differences in the heights of theplants or the aboveground and underground biomass between thetwo study sites (except for the density of the plants measured in a50 cm × 50 cm area) (see SI Table S1).

2.2. Gas sampling

At each site, four sampling plots (four replicates) were establishedalong a transect parallel with the river channel or the coastline at 3-mintervals. Enclosed, opaque static chambers were deployed for themea-surement of N2O emitted to the atmosphere. To ensure minimal physi-cal disturbance during chamber deployment and the measurementperiods, a wooden boardwalk was constructed to connect all plots.

The chambers had a polyvinyl chloride (PVC) bottom collar (35 cm× 35 cm × 30 cm in depth) and an opaque PVC upper chamber(35 cm× 35 cm× 140 cm in height). The bottom collarwas permanent-ly inserted into the marsh soil, with 2 cm left protruding above the soilsurface. In one side of the top chamber, a gas sampling port lined withrubber cushioning material was installed as an air-tight sealant 90 cmfrom the ground. The chamber was installed with an electric fan to en-sure slow but complete homogeneous air mixing. During the summerfieldmeasurement campaign, cotton quilts were used over the chambertops to ensure effective temperature control.

Monthly N2O measurements were made for a full year (November2013 to October 2014). Samples were collected between 8:00 am and10:00 am during low or a neap tides each month. Each flux measure-ment was performed for four samples collected at 10-min intervals.Samples of 50 mL were collected in 100-mL aluminum-foil gas samplebags (Dalian Delin Gas Packing Co., Ltd., China) using 60-mL gas-tight

ide fluxes and soil denitrification rates in subtropical freshwater and16/j.scitotenv.2017.10.175

Fig. 1. Location map of the two study sites in the Min River estuary.

3X. Wang et al. / Science of the Total Environment xxx (2017) xxx–xxx

syringes (Jiangsu Kangyou Medical Instrument Co., Ltd., China)equipped with three-way locks.

2.3. Estimation of N2O flux

N2O concentrations were determined using a gas chromatograph(GC-2014, Shimadzu, Japan) with an electron capture detector (ECD).The injection port, column, and detector temperatures were set at 50,60, and 320 °C, respectively, with helium as the carrier gas at a flowrate of 30 mL min−1. Rates of increasing N2O concentrations in thechambers were determined using linear regression, and data wererejected if the R2 value of the regression was ≤0.80 (Hirota et al.,2007). N2O fluxes were calculated using the following equation fromWang et al. (2015):

F ¼ dcdt

MV

273273þ T

� �H ð1Þ

where F is the N2O flux (μgm−2 h−1); dc/dt is the slope of the linear re-gression for thegasconcentrationgradientwith time(μmolmol−1h−1);M is the molecular mass of N2O (g mol−1); V is the N2O molar volume(L mol−1) under standard conditions; T is the air temperature duringsampling (°C); and H is the height of the measurement chamber (m).

2.4. Measurement of environmental variables

Various environmental factors were measured, including air tem-perature, soil temperature, soil electrical conductivity (EC), and soilpH. Air temperature was measured 1.5 m above ground with a Kestrel3500 hand-held weather meter (Kestrel-3500, USA) and recorded asan average of four consecutive measurements. Soil temperature andEC were measured using a 2265FS EC Meter (Spectrum TechnologiesInc., USA), and soil pHwasmeasured with an IQ150 instrument (IQ Sci-entific Instruments, USA) at the bottom collar at 5-cm depth. Besideeach static chamber, a 5-cm-deep soil sample was excavated and a 3-mL subsample was taken with a 5-mL syringe to determine the soilbulk density and moisture content.

Pore water sampling was done with 5-cm-external-diameter PVCpore water samplers installed at three different depths (5, 15, and25 cm) near each bottom collar (Ding et al., 2003; Strack and

Please cite this article as: Wang, X., et al., Seasonal variations of nitrous oxbrackish tidal marshes of..., Sci Total Environ (2017), https://doi.org/10.10

Waddington, 2008). The 5-cm sections of the samplers protrudedabove the soil surface and were sealed tightly with covers. All samplerswere emptied two days prior to gas sampling. Forty-milliliter porewater samples were collected when gas was sampled, and the physico-chemical properties of the pore water weremeasured in the laboratory:NH4

+-N and NO3−-N concentrations were determined through flow in-

jection analysis (FIA) using a Skalar San++ (Skalar Analytical B·V, Neth-erlands); Cl− and SO4

2− concentrations were determined via ionchromatography (IC) using a Dionex ICS-2100 integrated reagent-freesystem (Dionex, Sunnyvale, USA); and dissolved organic carbon (DOC)concentrationwas determinedwith the high-temperature (680 °C) cat-alytic oxidation method using a Shimadzu TOC-VCPH analyzer(Shimadzu, Japan).

2.5. Soil denitrification

2.5.1. Soil cores and overlying water collectionSoil cores were collected on the same days as the gas samples in No-

vember 2013, February 2014, May 2014, and August 2014. Two undis-turbed 15-cm-deep soil cores were collected using two 30-cm-longand 5-cm-inner-diameter PVC corers next to each bottom collar. Thebottoms of the corers were plugged tightly with PVC plugs. Besideeach soil core sampling site, a 15-cm-deep soil sample was extractedto determine the physicochemical properties of the soil. Overlyingwater samples were also collected for incubation and analysis.

2.5.2. Incubation and sampling in the laboratorySoil cores and overlying water samples were equilibrated for 12 h in

the laboratory to minimize the effects of disturbances during sampling,transportation, and incubation setup. Overlying water samples from therespective study sites were carefully transferred to the correspondingcoring tubes. After 2 h, eight soil cores fromeach study sitewere dividedinto two subgroups to conduct the inhibition experiments (Wang et al.,2007).

The first subgroup (A) was the “beginning cores”, which weresacrificed at the start of the inhibition experiment. Sixty milliliter sub-samples of the overlying water were drawn quickly and injected intothe 60-mL headspace vials, pre-equipped with 0.3 mL of 38% formalde-hyde, and the lidwas tightened. The second subgroup (B)was the “acet-ylene inhibition samples”. Ten percent of the overlying water in each

ide fluxes and soil denitrification rates in subtropical freshwater and16/j.scitotenv.2017.10.175

4 X. Wang et al. / Science of the Total Environment xxx (2017) xxx–xxx

tubewas replacedwith acetylene-saturatedwater, whichwas preparedfrom the overlying water by replenishing acetylene. Then, 0.2 mL ofacetylene-saturated water was injected into the soil through 1-mmholes (filled with silicone rubber) using microsyringes. The holes werepre-drilled at 2-cmvertical intervals in the bottom15 cmof the PVC cor-ing tubes. The tops of the tubes were tightly sealed upon acetyleneinjection.

Group B cores were put in a cyclotron oscillator (HY-5, JintanHongke, China) on mild concussion settings to ensure homogeneity ofthe overlying water and exchange of materials between the interfacewhile avoiding large disturbances. All incubations were done in thedark for 8 h. After incubation, 60 mL of the overlying water wasdrawn and quickly injected into 60-mL headspace vials, pre-equippedwith 0.3mL 38% formaldehyde. These vials were thenmixed and tightlyclosed. At the beginning and the end of each inhibition experiment, thetemperature and salinity of the extracted overlying water were mea-sured using the 2265FS EC Meter (Spectrum Technologies Inc., USA)and a Salt 6+ salinity meter (Oakton Instruments, IL, USA).

2.5.3. Estimation of dissolved N2O concentration in overlying waterThe dissolved N2O concentration in the overlying water was mea-

sured using static headspace-gas chromatography (Walter et al.,2005). Next, 30 mL of ultra-pure nitrogen was injected into each 60-mL headspace vial, and each vial was then filled with the incubatedsample of overlying water to exchange 30 mL of overlying water. The60-mL headspace vials were intensively shaken for 30 min to equili-brate the gas and the liquid phases. Fifteen milliliters of the headspacegas was sampled and injected into 50-mL aluminum foil gas samplebags using 30-mL gas-tight syringes equipped with three-way locks.N2O concentrations of the gas samples (Cg) were measured using thesame approach described in Section 2.3. The dissolved N2O concentra-tion in the overlying water was determined according to backward cal-culation from the headspace gas concentration using Henry's law (e.g.Musenze et al., 2014, 2015, 2016) and the constants of Weiss andPrice (1980), based on the overlying water temperature, water salinity,and the headspace gas N2O concentration (Eq. 2 and Eq. 3):

C ¼ CgK0 þ β=22:4 ð2Þ

lnK0 ¼ A1 þ A2 100=Tð Þ þ A3 ln T=100ð Þþ S B1 þ B2 T=100ð Þ þ B3 T=100ð Þ2

h ið3Þ

where C is the dissolved N2O concentration in the overlying water(μmol L−1); Cg is the N2O concentration in the air (μL L−1); K0 is thesolubility of N2O (mol L−1); β is the volume ratio of gas and liquid inthe headspace vial (β = 1 in this study); A1, A2, A3, B1, B2, and B3 areconstants (−62.7062, 97.3066, 24.1406, −0.05842, 0.033193, and−0.0051313, respectively); T is the thermodynamic temperature ofthe overlying water (K); and S is the salinity of the overlying water(‰) (Weiss and Price, 1980).

2.5.4. Estimation of soil denitrification rateThe soil denitrification rate was calculated using different N2O con-

centrations in overlying water between group A and group B (Eq. 3)(Yang et al., 2011):

RD ¼ CB−CAHw

t2 ð4Þ

where RD is the soil denitrification rate (μmol Nm−2 h−1); CA is theN2Oconcentration in the overlying water of group A (nmol L−1); CB is theN2O concentration in the overlying water of group B (nmol L−1); Hw

is the height of the overlyingwater in the PVC tube (0.15 m); t is the in-cubation time (h); and CA and CB are calculated using Eq. 2.

Please cite this article as: Wang, X., et al., Seasonal variations of nitrous oxbrackish tidal marshes of..., Sci Total Environ (2017), https://doi.org/10.10

2.5.5. Soil and overlying water characterizationTo determine the physicochemical properties of the soil, 15-cm-

deep soil samples were selected, air-dried, and ground using a ceramicmortar and pestle. Soil EC and pH were measured using a portable con-ductivity meter (2265FS EC Meter, Spectrum Technologies Inc., USA)and a portable Starter ST300 pH meter (Ohaus, USA), respectively. Thesoil-water mixture was prepared to a ratio of 2.5:1 (water:soil) and vi-brated for 30min. NH4

+-N and NO3−-N concentrations were determined

by measuring the soil leaching liquor (extracted with 2 mol L−1 KCl so-lution) usingflow injection analysis (FIA) (see Section 2.4). Total carbon(TC) and total nitrogen (TN) concentrations of soil samples weredetermined using a Vario MAX CN element analyzer (ElementarAnalysensysteme GmbH, Hanau, Germany).

The salinity of the overlying water was measured directly using aSalt 6+ salinitymeter (Oakton Instruments, IL, USA). Cl− and SO4

2− con-centrations in the overlying water were determined using the IC meth-od on 0.2-μm filters and syringe-filtered samples. NH4

+-N and NO3−-N

concentrations were measured using the FIA method described above,whereas DOCwas determined using the high-temperature catalytic ox-idation method (see Section 2.4).

2.6. Statistical analysis

The significance of the observed differences in N2O fluxes and envi-ronmental variables between the two sampling siteswasdetermined bycomparing the repeatedmeasurement values using analysis of variance(ANOVA). Multiple measurements at a given site over 12months repre-sented the repeated variables. The differences in denitrification ratesand properties of the soil and overlying water at each site in differentseasons were examined using least-significant difference (LSD) test inone-way ANOVA. Factors controlling the seasonal variation of N2Ofluxes and soil denitrification rates were examined with Pearson corre-lation analysis. All statistical analyses were performed using the SPSSstatistical package (SPSS 17.0 for Windows). The reported statisticswere interpreted to be inferential (Musenze et al., 2015).

3. Results

3.1. Monthly and seasonal variations of N2O fluxes

N2O fluxes from both sites showed distinct temporal variability(Fig. 2). In the Daoqingzhou wetland, the minimum and maximumN2O fluxes from the shichito matgrass marsh to the atmosphere were27.11 and 80.30 μg m−2 h−1 in January (winter) and June (summer),respectively. In the Shanyutan wetland, N2O fluxes ranged from 13.22to 41.54 μg m−2 h−1 throughout the year, as measured in November(autumn) and July (summer), respectively. Overall, the average N2Ofluxes were significantly higher (p b 0.05) in Daoqingzhou (48.81 ±9.01 μg m−2 h−1) than in Shanyutan (27.69 ± 4.01 μg m−2 h−1). Forboth sites, themaximum average N2O fluxes were observed in the sum-mer (Fig. 2b).

3.2. Environmental variables

Details of the monthly variations in air temperature, soil temper-ature, EC, pH, bulk density, and moisture content at our study sitesfor the same study period were reported previously (Hu et al.,2016). Briefly, air and soil temperatures in the Daoqingzhou andShanyutan wetlands demonstrated significant and similar monthlyvariations. There were no significant differences in air temperaturesbetween the twowetlands, but the average soil temperature was sig-nificantly higher (p b 0.05) in Daoqingzhou than in Shanyutan by0.78 °C. The soil EC in Daoqingzhou and Shanyutan varied between0.20 and 0.49 mS cm−1 and 0.80 and 6.62 mS cm−1, respectively.The average soil EC in Shanyutan (3.66 ± 0.84 mS cm−1) wassignificantly higher (p b 0.05) than that in Daoqingzhou (0.35 ±

ide fluxes and soil denitrification rates in subtropical freshwater and16/j.scitotenv.2017.10.175

Fig. 2.Monthly and seasonal variations of N2O flux in Daoqingzhou and Shanyutan. Barsshow the averages of four months of measurements. Error bars are the standard errorson the mean values (n = 4). “Winter” includes N2O fluxes measured in December 2013,January 2014, and February 2014; “Spring”: March, April, and May 2014; “Summer”:June, July, and August 2014; and “Autumn”: September 2014, October 2014, andNovember 2013. Different lowercase letters indicate significant differences betweenseasons at the same sampling site; different capital letters indicate significantdifferences between sampling sites during the same season. The results of the repeatedmeasurements ANOVA are also reported; asterisks show the significance of factors (site(S) and month (M)) and their interactions: **p b 0.01, *p b 0.05, ns = p N 0.05.

Fig. 3.Monthly variations of porewater concentrations of NH4+-N (a), NO3

−-N (b), Cl− (c), SO42−

three sampled depths. Error bars are the standard errors of the average values; n=4. The resultthe factors (site (S) and month (M)) and their interactions: ***p b 0.001, **p b 0.01, *p b 0.05, n

5X. Wang et al. / Science of the Total Environment xxx (2017) xxx–xxx

Please cite this article as: Wang, X., et al., Seasonal variations of nitrous oxbrackish tidal marshes of..., Sci Total Environ (2017), https://doi.org/10.10

0.04 mS cm−1). The soil pH in Daoqingzhou and Shanyutan variedbetween 4.81 and 6.97 and 5.25 and 7.87, respectively. The averagesoil pH in Daoqingzhou (5.80 ± 0.41) was significantly lower (p b

0.05) than that in Shanyutan (6.60± 0.37). There were no significantdifferences in the bulk density or moisture content of the soil be-tween the two sites.

Monthly variations of pore water concentrations of NH4+-N (Fig. 3a),

NO3−-N (Fig. 3b), Cl− (Fig. 3c), SO4

2− (Fig. 3d), and DOC (Fig. 3e), interms of the mean values at three sampling depths (5, 15, and 25 cm)in theDaoqingzhou and Shanyutanwetlands are shown in Fig. 3. No sig-nificant differences were found between the two sites in NH4

+-N andNO3

−-N concentrations in pore water. However, Cl− and SO42− concen-

trations in pore water in Shanyutan were 111 and 24 times greaterthan the respective concentrations in Daoqingzhou. The two studysites exhibited similar monthly trends for Cl− concentrations. The aver-age DOC concentration in the pore water was significantly lower (p b

0.05) in Daoqingzhou than in Shanyutan for most of the year.

3.3. Influence of environmental variables on N2O fluxes

Therewere strong positive correlations between theN2Oflux and airtemperature/soil temperature at both sites (see Table S2). The soil ECshowed a strong positive correlation with N2O flux in Daoqingzhou,but a negative correlation with N2O flux in Shanyutan. The soil pH andDOC concentrations of the pore water showed strong negativecorrelations with N2O flux in Daoqingzhou, whereas greater NH4

+-Nconcentration promoted N2O flux significantly. Negative correlation be-tween pore water SO4

2− concentration and N2O flux was only detectedin Shanyutan.

(d), and DOC (e) in Daoqingzhou and Shanyutan. Data represent the average values of thes of the repeatedmeasurements ANOVA are also reported; asterisks indicate significance ofs = p N 0.05.

ide fluxes and soil denitrification rates in subtropical freshwater and16/j.scitotenv.2017.10.175

Fig. 4. Seasonal variations in soil denitrification rates in Daoqingzhou and Shanyutan. Barsrepresent the averages of monthly values, whereas the error bars are the standard errorsof the average monthly values (n = 3). Different lowercase letters indicate significantdifferences between seasons at the same sampling site; different capital letters indicatesignificant differences between different sampling sites during the same season.

6 X. Wang et al. / Science of the Total Environment xxx (2017) xxx–xxx

3.4. Seasonal variations of soil denitrification rates

Significant seasonal variations in soil denitrification rates were ob-served in the Daoqingzhou and Shanyutan wetlands (Fig. 4). InDaoqingzhou, the maximum and minimum soil denitrification rateswere 80.91 ± 19.57 and 3.37 ± 1.18 μmol N m−2 h−1 in summer andautumn, respectively. In Shanyutan, the soil denitrification rates rangedfrom 0.33± 0.14 to 11.46± 1.64 μmol Nm−2 h−1, with the highest de-nitrification rates detected in spring. Overall, denitrification rates weresignificantly higher in Daoqingzhou (p b 0.05) than in Shanyutanthroughout the study period.

3.5. Properties of soil and overlying water

There were no significant differences in soil pH between the four sea-sons in Daoqingzhou (Table 1). However, other soil physicochemicalproperties demonstrated significant seasonal variations in two wetlands.The seasonal coefficients of variability for EC, NH4

+-N, and NO3−-N

concentrations were relatively higher than the other physicochemicalproperties in both wetlands. The EC, pH, and NO3

−-N concentrations inthe soil were significantly higher (p b 0.05) in Shanyutan than inDaoqingzhou. In contrast, the NH4

+-N concentrations, TC concentrations,

Table 1Seasonal variations in soil physicochemical properties in the Daoqingzhou and Shanyutan wetlconcentrations are expressed per kg of dry soil).

Site Season EC pH NH4+-N

(mS cm−1) (mg kg−1)

Daoqingzhou Autumn 0.26 ± 0.04a B 5.02 ± 0.19a B 38.89 ± 4.06b A

Winter 0.25 ± 0.03a B 5.07 ± 0.20a B 42.06 ± 6.20ab A

Spring 0.31 ± 0.05a B 4.85 ± 0.08a B 41.57 ± 5.51ab A

Summer 0.16 ± 0.02b B 4.96 ± 0.09a B 50.04 ± 5.65a A

CV 25.87% 1.85% 11.14%Shanyutan Autumn 3.13 ± 0.51a A 6.54 ± 0.14c A 21.20 ± 2.29b B

Winter 2.75 ± 0.61a A 6.65 ± 0.15bc A 26.98 ± 4.86ab B

Spring 1.61 ± 0.48b A 6.89 ± 0.11a A 30.28 ± 4.17a B

Summer 1.14 ± 0.22b A 6.75 ± 0.12ab A 26.06 ± 6.72ab B

CV 43.38% 2.23% 14.38%

Different lowercase letters indicate significant differences between seasons at the same samplisites during the same season.

Please cite this article as: Wang, X., et al., Seasonal variations of nitrous oxbrackish tidal marshes of..., Sci Total Environ (2017), https://doi.org/10.10

and carbon-to-nitrogen ratios (C/N) of the soil were higher (p b 0.05) inthe Daoqingzhou wetland.

There were significant seasonal variations in the physicochemicalproperties of the overlying water at both sites (Table 2). Seasonal vari-ability in all physicochemical properties of the overlying water, exceptfor DOC concentration, was higher (p b 0.05) in the Shanyutanwetland.The salinity and concentrations of Cl−, SO4

2−, NH4+-N, and DOC of the

overlying water were significantly higher (p b 0.05) in Shanyutan thanin Daoqingzhou. However, NO3

−-N concentrations were higher (p b

0.05) in the Daoqingzhou wetland.

3.6. Influence of soil and overlying water on soil denitrification

For incubation, the overlyingwater temperatures in autumn, winter,spring, and summer were 24.0± 0.1, 18.4± 0.5, 28.0 ± 0.2, and 30.5±0.1 °C, respectively. Linear correlation models were fitted to the rela-tionship between the overlyingwater temperatures and the soil denitri-fication rates in bothwetlands. There was a significant linear increase inthe soil denitrification rate with increased overlying water temperaturein both wetlands (n = 16, p b 0.01) (Fig. 5).

The relationship between soil physicochemical properties anddenitrification rates in the Daoqingzhou and Shanyutan wetlands wasinconsistent (Table S3). In the Daoqingzhou wetland, the soil denitrifi-cation rate showed negative correlations with EC and C/N and a positivecorrelation with NH4

+-N concentration. In contrast, in the Shanyutanwetland, a negative (and positive) correlation was detected betweenthe denitrification rate and EC (and TN).

The relationship between overlying water physicochemical proper-ties and denitrification rates in the Daoqingzhou and Shanyutanwetlands is shown in Table S3. The concentrations of Cl− and SO4

2−

(NH4+-N) in the overlying water had significantly negative (positive)

correlations with soil denitrification rates in both wetlands. However,a strong negative correlation between the denitrification rate and thesalinity and DOC concentration (NO3

−-N concentration) of the overlyingwater was detected only in Shanyutan (Daoqingzhou).

4. Discussion

4.1. Nitrous oxide fluxes in estuarine tidal marshes

Both the fresh- and brackish tidal marshes of the Min River estuarywere found to be strong sources of atmospheric N2O. Similar resultshave been previously reported from different climatic zones (Table 3).Estuarine tidal marshes, which are gateways of terrestrial nitrogen dis-charge into oceans, provide favorable conditions for the production andsubsequent emission of N2O, including availability of organic carbonand nitrogen, and alternating dry and wet environments (Galloway etal., 2003; Moseman-Valtierra et al., 2011; Wang et al., 2016). Studies

ands of the Min River estuary (numbers are “averages ± standard errors”; n=4. Nutrient

NO3−-N TC TN C/N

(mg kg−1) (g kg−1) (g kg−1)

0.64 ± 0.20a A 21.48 ± 0.89a A 1.79 ± 0.07bc A 12.01 ± 0.20a A

0.16 ± 0.05b B 18.88 ± 0.69c A 1.71 ± 0.06c B 11.07 ± 0.13b A

0.17 ± 0.03b B 20.92 ± 1.28ab A 1.90 ± 0.12ab A 11.05 ± 0.21b A

0.31 ± 0.18ab A 19.84 ± 0.81bc A 1.96 ± 0.10a A 10.14 ± 0.17c A

70.15% 5.69% 6.14% 6.87%0.52 ± 0.14a A 18.70 ± 0.96ab B 1.83 ± 0.07b A 10.21 ± 0.23a B

0.28 ± 0.03bc A 17.61 ± 0.67b B 1.84 ± 0.06b A 9.60 ± 0.27b B

0.26 ± 0.04c A 19.03 ± 0.69a B 2.00 ± 0.07a A 9.52 ± 0.24b B

0.40 ± 0.11b A 18.27 ± 1.06ab A 2.08 ± 0.10a A 8.78 ± 0.17c B

32.92% 3.33% 6.39% 6.17%

ng site; different capital letters indicate significant differences between different sampling

ide fluxes and soil denitrification rates in subtropical freshwater and16/j.scitotenv.2017.10.175

Table 2Seasonal variations in thephysicochemical properties of overlyingwater in theDaoqingzhou and Shanyutanwetlands of theMin River estuary (numbers are “averages± standard errors”;n = 4).

Site Season Salinity (‰) Cl− (mg L−1) SO42− (mg L−1) NH4

+-N (mg L−1) NO3−-N (mg L−1) DOC (mg L−1)

Daoqingzhou Autumn 0.21 ± 0.00a B 13.02 ± 0.51a B 17.63 ± 0.06a B 0.05 ± 0.00b B 1.87 ± 0.01a A 10.68 ± 0.18c B

Winter 0.21 ± 0.00a B 9.21 ± 0.09b B 16.19 ± 0.02b B 0.19 ± 0.02b B 1.65 ± 0.03b A 23.86 ± 0.83b B

Spring 0.17 ± 0.01c B 2.84 ± 0.20d B 9.41 ± 0.08c B 0.23 ± 0.02ab B 0.88 ± 0.01c A 28.47 ± 2.91a A

Summer 0.19 ± 0.00b B 4.28 ± 0.25c B 8.43 ± 0.10d B 0.24 ± 0.01a B 0.82 ± 0.02c A 4.64 ± 0.13d B

CV 10.64% 63.63% 36.13% 48.34% 40.91% 65.78%Shanyutan Autumn 8.10 ± 0.00a A 5659.08 ± 40.05a A 771.49 ± 7.58a A 0.80 ± 0.01b A 0.72 ± 0.01a B 22.79 ± 0.07b A

Winter 6.85 ± 0.03b A 4358.99 ± 4.43b A 626.49 ± 1.04b A 0.28 ± 0.03c A 0.51 ± 0.01c B 34.64 ± 3.87a A

Spring 1.14 ± 0.00c A 513.27 ± 1.18c A 106.95 ± 0.75c A 1.28 ± 0.02a A 0.66 ± 0.01b B 12.71 ± 0.29c B

Summer 1.03 ± 0.00d A 396.19 ± 5.46d A 27.43 ± 0.78d A 0.83 ± 0.01b A 0.08 ± 0.00d B 7.59 ± 0.05d A

CV 86.96% 98.21% 96.84% 50.95% 58.96% 61.46%

Different lowercase letters indicate significant differences between seasons at the same sampling site; different capital letters indicate significant differences between different samplingsites during the same season.

7X. Wang et al. / Science of the Total Environment xxx (2017) xxx–xxx

on N2O sources/sinks and their key controls are necessary for (1) accu-rate compilation of N2O emission inventories of tidal marshes, (2) morereliable prediction of N2O emissions in changing climates and environ-ments in the future and (3) better accountability of the global N2Obudget.

Fluxes of N2Owere found to be significantly higher in the freshwatermarsh than in the brackishmarsh. Although these findings are in agree-ment with previous studies (Welti et al., 2016; Smith et al., 1983), re-ported data on the effect of salinity on N2O emissions in estuarinemarshes are largely inconsistent (Table 4). The interactions betweenthe drivers of N2O production in different environments are very com-plex. In the investigated estuarine tidalmarshes, the key environmentalfactors included nitrogen availability, tidal water level, and rooted veg-etation depth (Hernandez andMitsch, 2006;Wang et al., 2007). The re-sults of this study indicate that N2O emissions were higher in thefreshwater marsh than in the brackish (oligohaline) marsh (Allen etal., 2011; Chen et al., 2010), even though the range of salinity over thestudy period was narrow. In general, N2O emission studies in subtropi-cal estuarine areas have shown similar trends (e.g. Musenze et al., 2014;Sturm et al., 2016). Previously, it has been argued that salinity could in-hibit N2O production in many ways, including through physical, chem-ical, and physiological processes (Edmonds et al., 2009; Giblin et al.,2010). This study was not designed to decouple the contributions ofthe different N2O production processes andmechanisms aside from de-nitrification rates; however, the results clearly reveal that increased in-trusion of saline seawaters into estuarine freshwater wetlands couldreduce N2O emissions. This phenomenon would lead to negative feed-back on sea-level rise.

4.2. Seasonal variations of nitrous oxide fluxes in estuarine tidal marshes

There are various controls on N2O emissions from aquatic systems.Air and soil temperature are key drivers of the temporal dynamics of

Fig. 5. Relationship between soil denitrification rate and tem

Please cite this article as: Wang, X., et al., Seasonal variations of nitrous oxbrackish tidal marshes of..., Sci Total Environ (2017), https://doi.org/10.10

N2O flux. Soil temperature controls the decomposition rate of soil or-ganic matter through its effect on microbial activity, enzymes, and me-tabolism, and consequently regulates the rates of N2O production,consumption, and ultimately emission into the atmosphere. Tempera-ture would also affect the physical properties of soil, as well as the solu-bility of N2Owithin both the porewater and the overlyingwater (Weissand Price, 1980;Wanninkhof, 1992). Temperature would also affect thewater-air gas transfer rate (Guo et al., 2013;Musenze et al., 2014, 2015;Zhu et al., 2014). The significant positive correlation between N2O fluxand temperature (Table S2) indicates that temperature is an importantcontrol of the seasonal variability of N2O flux. The estimated annual N2Ofluxes were 429.6 mg N2O m−2 y−1 and 244.1 mg N2O m−2 y−1 in theDaoqingzhou and Shanyutan wetlands, respectively. During summer,with elevated temperatures (measured average air (soil) temperatureswere 35.3 (29.5)°C and 32.0 (29.1)°C in the Daoqingzhou andShanyutan wetlands, respectively), the summer fluxes accounted for~33% and ~34% of the annual fluxes in the Daoqingzhou and Shanyutanwetlands, respectively. This finding confirms that N2O emissions in sub-tropical aquatic systems are normally elevated during the summermonths relative to other seasons (Musenze et al., 2014, 2015; Allen etal., 2011). N2O fluxes also increase with sediment and pore water tem-peratures (Zheng et al., 1997; Li et al., 2010). Zheng et al. (1997) showedthat 67% of N2O emissionswere released at temperatures of between 15and 25 °C. Li et al. (2010) showed that temperatures of 25–36 °C weresuitable for microbial N2O production (91.3% of total emissions) in aPhragmites australis ecosystem on eastern Chongming Island. However,the observation of elevated emissions in the summer should beinterpreted cautiously. Subtropical summers are hot and wet, with in-creased catchment flows, organic matter, and terrigenous nitrogen de-posits, which could sustain increased greenhouse gas production ratesfor longer periods of time (Musenze et al., 2014, 2015, 2016). These fac-tors could be the reason that aside from elevated summer emissions,seasonal fluxes were similar.

perature in the Daoqingzhou and Shanyutan wetlands.

ide fluxes and soil denitrification rates in subtropical freshwater and16/j.scitotenv.2017.10.175

Table 3N2O fluxes from estuarine freshwater and brackish tidal marshes in different climate zones worldwide.

Location Climate zone Vegetation Habitat N2O flux (μg m−2 h−1) References

Min River estuary, China Subtropical Cyperus malaccensis Freshwater 48.81 ± 9.01 This studySavannah River, USA Subtropical Taxodium dstichum;

Nyssa aquaticaFreshwater −3.32 ± 5.79 Krauss and Whitbeck, 2012

Cooroon Cooroonphah Creek, Australia Subtropical Gahnia sieberiana;Empodisma minus;Gleichenia spp

Freshwater 59.0 ± 24.3 Welti et al., 2016

Dovey estuary, UK Temperate Agrostis stolonifera Freshwater 20.2 ± 5.5 Dausse et al., 2012Min River estuary, China Subtropical Cyperus malaccensis Brackish 27.69 ± 4.01 This studySavannah River, USA Subtropical Taxodium dstichum;

Nyssa aquaticaBrackishBrackish

−1.91 ± 4.058.91 ± 4.77

Krauss and Whitbeck, 2012

Futian, Deep Bay region, China Tropical Kandelia obovata Brackish 757.24 Chen et al., 2010Mai Po, Deep Bay region, China Tropical Kandelia obovata Brackish 32.1–533.7 Chen et al., 2012Yellow River estuary, China Temperate Suaeda salsa Brackish 5.2–15.2 Sun et al., 2017Dovey estuary, UK Temperate Festuca rubra Brackish 4.6 ± 2.5 Dausse et al., 2012Nakaumi lagoon, Japan Temperate Carex rugulosa;

Phragmites australis;Solidago altissima

Brackish −30–1 Hirota et al., 2007

8 X. Wang et al. / Science of the Total Environment xxx (2017) xxx–xxx

4.3. Soil denitrification rates in estuarine tidal marshes

There was strong seasonal variability in sediment denitrificationrates in the estuarine tidal marshes. The highest denitrification rates oc-curred during periods of elevated temperatures in the summer andspring in the Daoqingzhou and Shanyutan wetlands, respectively,whereas the lowest denitrification rates were observed in autumn andwinter. The relationship between denitrification rate and overlyingwater temperaturewas strong and positive (Fig. 5), with the denitrifica-tion rates increasing between the temperature range of ~18–30 °C inboth tidal marshes. Recently, the optimal temperature for N2O produc-tion via the denitrification pathway has been reported as 25 °C in theinter-tidal zone of the Yellow River estuary, China (Sun et al., 2014).Wang et al. (2007) also found temperature to be a key control of sedi-ment denitrification rates in the tidal flat of the Changjiang estuary.The influence of temperature on denitrification rate is likely becauseof its impact on microbial activity and reaction rates. Under relativelyhigh-temperature conditions, strong microbial activity in sedimentsleads to higher denitrification potential and higher mineralization.This study found no significant relationship between soil NO3

−-N con-tent and the denitrification rate. Although this finding could easily beinterpreted as indicating that NO3

−-N concentrations are not amain fac-tor affecting the denitrification rates at the study sites, it could also be anindication that there was an ample supply of NO3

−-N to meet the nitro-gen demands of the denitrifyingbacteria, or simply the result of lowvar-iability of NO3

−-N concentration at the study sites over the study period.Therefore, this thrust the other drivers (e.g., temperature) into key rolescontrolling N2O production via the denitrification pathway.

4.4. Effects of salinity on estuarine soil denitrification rates

The results of this study show that the average soil denitrificationrates were significantly higher in the freshwater marsh of theDaoqingzhouwetland than in the brackishmarsh of the Shanyutanwet-land (Fig. 4), although NO3

−-N concentrations in both wetlands weresimilar. Salinity (together with EC and Cl− concentrations) of the soil

Table 4Comparison of N2O emissions from various estuarine tidal marshes with different salinity leve

Location Salinity (‰)

Barataria Basin, USA Fresh①; Brackish②; Saline③Dovey estuary, England 1.3①;2.6②; 9.6③; 17.3④Savannah River, USA b0.2①; 0.4–2.1②; 1.9–6.8③Cooroon Cooroonphah Creek, Australia Freshwater①; Oligohaline saltwaMin River estuary, China 0.20①; 3.79②

“≈” indicates that there is no significant difference between two sites.

Please cite this article as: Wang, X., et al., Seasonal variations of nitrous oxbrackish tidal marshes of..., Sci Total Environ (2017), https://doi.org/10.10

and the overlying water were significantly higher in Shanyutan thanin Daoqingzhou (Table 1 and Table 2) and had significant negativecorrelations with soil denitrification rate (Table S3). With all the otherpresumed critical controls of denitrification and N2O production inves-tigated found to be similar in both wetlands, and only salinity and itsprecursors found to be significantly higher in Shanyutan than inDaoqingzhou (Table 1 and Table 2), this study demonstrates that salin-ity was an important controlling parameter responsible for the lowersoil denitrification rates in the brackish marsh (Fear et al., 2005;Zimmerman and Benner, 1994; Giblin et al., 2010; Musenze et al.,2014). Salinity values below 6 psu favor retention of NH4

+ in sedimentand higher denitrification rates (Zimmerman and Benner, 1994). Thisphenomenon is likely the reason for the significant relationship be-tween sediment NH4

+-N concentration and denitrification rate inDaoqingzhou, whereas no significant relationship was detected inShanyutan. Salinity also negatively impacts microbial diversity(Francis et al., 2003;Mosier and Francis, 2008;Moin et al., 2009) and ac-tivity (Wang et al., 2010), especially of ammonia-oxidizing bacteria(AOBs). Nonetheless, the influence of salinity is still poorly understood,and conflicting reports suggest that salinity has no influence on denitri-fication rate (Wang et al., 2007; Fear et al., 2005). Therefore, furtherstudies are still needed for a better understanding of the mechanismsand impacts of salinity on biogenic N2O-producing processes(Musenze et al., 2014, 2015).

4.5. Effects of other environmental factors on denitrification rates and ni-trous oxide fluxes

Denitrification rates had a strong negative correlation with SO42−

concentrations in both the Daoqingzhou and Shanyutan wetlands(Table S3). This relationship was stronger in Shanyutan than inDaoqingzhou, likely because of the significantly higher SO4

2− concentra-tions in the Shanyutan wetland. SO4

2− concentrations in freshwaterbodies are typically lower than those in saline/sea water (Musenze etal., 2014; Musenze et al., 2016). The higher SO4

2− concentrations, thesignificant negative relationship between SO4

2− and denitrificationrates, and low N2O saturation levels in the Shanyutan wetland are

ls.

N2O flux References

① N ② N ③ Smith et al., 1983① b ② ≈ ③ ≈ ④ Dausse et al., 2012① ≈ ② ≈ ③ Krauss and Whitbeck, 2012

ter② ① N ② Welti et al., 2016① N ② This study

ide fluxes and soil denitrification rates in subtropical freshwater and16/j.scitotenv.2017.10.175

9X. Wang et al. / Science of the Total Environment xxx (2017) xxx–xxx

strong indicators of the negative effect of pore water SO42− concentra-

tions on N2O production and subsequent emissions. Under strong re-ducing conditions, SO4

2− can be reduced to H2S. The H2S thusproduced can influence the activities of both nitrifiers and denitrifiersand their effects on N2O production (Giblin et al., 2010). The reductionof SO4

2− can also enhance dissimilatory nitrate reduction to ammonium(DNRA) (Tiedje, 1988) while inhibiting denitrification (Senga et al.,2006; Larsen et al., 2010), especially at higher metal sulfide concentra-tions (Brunet andGarcia-Gil, 1996),which directly impactsN2O produc-tion and flux. The strong negative relationship of denitrification rateswith such SO4

2− concentrations in overlying water in Daoqingzhou sug-gests high sensitivity of denitrification to sulfates. Increased sulfateavailability in freshwater bodies originating from agricultural croplandsand other land uses, together with seawater intrusion into inland fresh-water bodies, could significantly affect nitrogen removal processes andN2O emissions. Therefore, in terms of climate feedback, sea level rise,which would lead to increased seawater intrusion into estuarine fresh-water tidalwetlands, is likely to lead to a reduction inN2O production inestuarine tidal wetlands.

During denitrification, denitrifying bacteria use nitrates as electronacceptors in the absence of oxygen for their respiration (respiratory de-nitrification). In the process of nitrate reduction, N2O is normally pro-duced as an intermediate that may either be emitted or furtherreduced under strong reducing conditions to the end product, N2 gas.Therefore, denitrification is strongly dependent on nitrate concentra-tion, and inmost cases, nitrates would be a limiting factor of denitrifica-tion (Moseman-Valtierra et al., 2011). In contrast, our data indicate thatsediment nitrate concentrations had no significant relationshipwith de-nitrification rates, but were strongly correlated with the NH4

+ concen-trations of overlying water. In a well-oxygenated water column,denitrification mainly occurs in the sediment (Musenze et al., 2014;Musenze et al., 2015). The overlyingwater columnwould be dominatedby nitrification because of the availability of both oxygen and NH4

+. It islikely thatmost nitrification occurs in the uppermost few centimeters ofthe sediments under reducing conditions because of the sensitivity ofnitrifiers to oxygen (Knowles, 1982). Therefore, the availability ofNH4

+ is important for this step. This process is the main source of NO3−

for denitrifiers (Seitzinger, 1988; LaMontagne et al., 2003), and is there-fore the reason for their low dependence on sediment NO3

− levels. Themost conducive conditions for biogenic greenhouse gas production areat the oxycline (Hynes and Knowles, 1984; Pina-Ochoa andÁlvarez-Cobelas, 2006; Musenze et al., 2016), where NO3

− would bereadily available for denitrifiers. These conditions would be synony-mous with coupled nitrification and denitrification, which is often mis-taken for the nitrifier-denitrifier pathway, the N2O production potentialof which remains poorly understood (Webster and Hopkins, 1996;Wrage et al., 2001).

Soil organic carbon, an electron donor and energy source for denitri-fiers, could significantly affect soil denitrification rates. In general, soil C/N is used as an indicator of soil organic carbon availability. Lower C/N in-dicates a higher decomposition rate of soil organic carbon (Guo et al.,2012). The decomposition of organic matter consumes a large amountof oxygen, which creates an anaerobic environment favorable for deni-trification. In this study, soil denitrification rates showed significantnegative correlations with C/N (Table S3), which suggests that high or-ganic carbon availability exerts a strong oxygen demand that wouldleave insufficient oxygen for nitrification and eventually support deni-trification. Alternatively, the high organic carbon availability in relationto the nitrogen levels could trigger DNRA (Liu et al., 2016), especially inthe Shanyutan wetland, since DNRA is favored by the labile C/NO3

−

levels (Silver et al., 2001).

5. Conclusions

Both freshwater and brackish tidal marshes of theMin River estuaryare strong emitters of atmospheric N2O with significant temporal

Please cite this article as: Wang, X., et al., Seasonal variations of nitrous oxbrackish tidal marshes of..., Sci Total Environ (2017), https://doi.org/10.10

variability. N2O emissions were significantly higher in the hot and wetsummer season and lower in the cold autumn and winter seasons.N2O emissions from the freshwater marsh were significantly higherthan those from the brackish marsh. Soil denitrification rates in thefreshwater and brackish tidal marshes showed strong seasonal variabil-ity; the maximum (minimum) values were observed in summer (au-tumn) and spring (winter), respectively. Denitrification was stronglydependent on soil temperature. Soil denitrification rates in the brackishmarshwere significantly lower than those in the freshwatermarsh. N2Ofluxes and soil denitrification rates demonstrated negative correlationwith salinity and sulfate (SO4

2−) concentrations in both the freshwaterand brackish tidal marshes. Increased intrusion of saline seawater intosubtropical tidal estuarine freshwater aquatic systems (e.g., wetlands)could reduce N2O emissions, and exert negative feedback with respectto climate change.

Acknowledgments

This study was supported by the National Natural Science Founda-tion of China (41371127), the Program for Innovative Research Teamsof Fujian Normal University (IRTL1205), and the Key Sciences and Tech-nology Project of Fujian Province (2014R1034-1). We thank WenlongZhang, Ping Yang, Weiqi Wang and Jiafang Huang for their help withfield sampling.

Appendix A. Supplementary data

Supplementary data to this article can be found online at https://doi.org/10.1016/j.scitotenv.2017.10.175.

References

Allen, D., Dalal, R.C., Rennenberg, H., Schmidt, S., 2011. Seasonal variation in nitrous oxideand methane emissions from subtropical estuary and coastal mangrove sediments,Australia. Plant Biol. 13 (1), 126–133.

Bauza, J.F., Morell, J.M., Corredor, J.E., 2002. Biogeochemistry of nitrous oxide productionin the red mangrove (Rhizophora mangle) forest sediments. Estuar. Coast. Shelf Sci.55 (5), 697–704.

Brunet, R.C., Garcia-Gil, L.J., 1996. Sulfide-induced dissimilatory nitrate reduction to am-monia in anaerobic freshwater sediments. FEMS Microbiol. Ecol. 21, 131–138.

Carson, M., Köhl, A., Stammer, D., Slangen, A.B.A., Katsman, C.A., van de Wal, R.S.W.,Church, J., White, N., 2016. Coastal sea level changes, observed and projected duringthe 20th and 21st century. Clim. Chang. 134, 269–281.

Chauhan, R., Datta, A., Ramanathan, A.L., Adhya, T.K., 2015. Factors influencing spatio-temporal variation of methane and nitrous oxide emission from a tropical mangroveof eastern coast of India. Atmos. Environ. 107, 95–106.

Chen, G.C., Tam, N.F.Y., Ye, Y., 2010. Summer fluxes of atmospheric greenhouse gases N2O,CH4 and CO2 from mangrove soil in South China. Sci. Total Environ. 408, 2761–2767.

Chen, G.C., Tam, N.F.Y., Ye, Y., 2012. Spatial and seasonal variations of atmospheric N2Oand CO2 fluxes from a subtropical mangrove swamp and their relationships withsoil characteristics. Soil Biol. Biochem. 48, 175–181.

Dausse, A., Garbutt, A., Norman, L., Papadimitriou, S., Jones, L.M., Robins, P.E., Thomas,D.N., 2012. Biogeochemical functioning of grazed estuarine tidal marshes along a sa-linity gradient. Estuar. Coast. Shelf Sci. 100, 83–92.

Denman, K.L., Brasseur, G., Chidthaisong, A., Ciais, P., Cox, P.M., Dickinson, R.E.,Hauglustaine, D., Heinze, C., Holland, E., Jacob, D., Lohmann, U., Ramachandran, S.,da Silva Dias, P.L., Wofsy, S.C., Zhang, X., 2007. Couplings between changes in the cli-mate system and biogeochemistry. In: Solomon, S., Qin, D., Manning, M., Chen, Z.,Marquis, M., Averyt, K.B., Tignor, M., Miller, H.L. (Eds.), Climate Change 2007: thePhysical Science Basis. Contribution ofWorking Group I to the Fourth Assessment Re-port of the Intergovernmental Panel on Climate Change. Cambridge University Press,Cambridge, United Kingdom and New York, NY, USA.

Ding,W.X., Cai, Z.C., Tsuruta, H., Li, X., 2003. Key factors affecting spatial variation ofmeth-ane emissions from freshwater marshes. Chemosphere 51, 167–173.

Edmonds, J.W., Weston, N.B., Joye, S.B., Mou, X., Moran, M.A., 2009. Microbial communityresponse to seawater amendment in low-salinity tidal sediments. Microb. Ecol. 58,558–568.

Fear, J.M., Thompson, S.P., Gallo, T.E., Pearl, H.W., 2005. Denitrification rates measuredalong a salinity gradient in the eutrophic Neuse River Estuary, North Carolina, USA.Estuaries 28 (4), 608–619.

Francis, C.A., O'Mullan, G.D., Ward, B.B., 2003. Diversity of ammonia monooxygenase(amoA) genes across environmental gradients in Chesapeake Bay sediments.Geobiology 1, 129–140.

Franklin, R.B., Morrissey, E.M., Morina, J.C., 2016. Changes in abundance and communitystructure of nitrate-reducing bacteria along a salinity gradient in tidal wetlands.Pedobiol. J. Soil Biol. https://doi.org/10.1016/j.pedobi.2016.12.002.

ide fluxes and soil denitrification rates in subtropical freshwater and16/j.scitotenv.2017.10.175

10 X. Wang et al. / Science of the Total Environment xxx (2017) xxx–xxx

Galloway, J.N., Aber, J.D., Erisman, J.W., Seitzinger, S.P., Howarth, R.W., Cowling, E.B.,Cosby, B.J., 2003. The nitrogen cascade. Bioscience 53 (4), 341–356.

Giblin, A.E., Weston, N.B., Banta, G.T., Tucker, J., Hopkinson, C.S., 2010. The effects of salin-ity on nitrogen losses from an oligohaline estuarine sediment. Estuar. Coasts 33,1054–1068.

Guo, R., Li, G., Jiang, T., Schuchardt, F., Chen, T., Zhao, Y., Shen, Y., 2012. Effect of aerationrate, C/N ratio and moisture content on the stability and maturity of compost.Bioresour. Technol. 112, 171–178.

Guo, N., Zhang, J., Kong, Q., Miao, M., Tian, L., 2013. Effect of temperature onammoxidation and nitrous oxide emission in partial nitrification (in Chinese). Chi-nese J. Environ. Eng. 7 (4), 1308–1312.

Herbert, E.R., Boon, P., Burgin, A.J., Neubauer, S.C., Franklin, R.B., Ardon,M., Hopfensperger,K.N., Lamers, L.P.M., Gell, P., 2015. A global perspective on wetland salinization: eco-logical consequences of a growing threat to freshwater wetlands. Ecosphere 6 (10),1–43.

Hernandez, M.E., Mitsch, W.J., 2006. Influence of hydrologic pulses, flooding frequency,and vegetation on nitrous oxide emissions from created riparian marshes. Wetlands26 (3), 862–877.

Hirota, M., Senga, Y., Seike, Y., Nohara, S., Kunii, H., 2007. Fluxes of carbon dioxide, meth-ane and nitrous oxide in two contrastive fringing zones of coastal lagoon, LakeNakaumi, Japan. Chemosphere 68, 597–603.

Hu, M.J., Ren, H.C., Ren, P., Li, J.B., Wilson, B.J., Tong, C., 2016. Response of gaseous carbonemissions to low-level salinity increase in tidal marsh ecosystem of the Min River es-tuary, southeastern China. J. Environ. Sci. https://doi.org/10.1016/j.jes.2016.05.009.

Hynes, R.K., Knowles, R., 1984. Production of nitrous oxide by Nitrosomonas europaea, ef-fects of acetylene, pH, and oxygen. Can. J. Microbiol. 30 (11), 1397–1404.

Jordan, S.J., Stoffer, J., Nestlerode, J.A., 2011. Wetlands as sinks for reactive nitrogen at con-tinental and global scales: a meta-analysis. Ecosystems 14, 144–155.

Knowles, R., 1982. Denitrification. Microbiol. Rev. 46 (1), 43–70.Knowles, N., 2002. Natural and management influences on freshwater inflows and salin-

ity in the San Francisco Estuary at monthly to interannual scales. Water Resour. Res.38 (12) (25-1-25-11).

Krauss, K.W., Whitbeck, J.L., 2012. Soil greenhouse gas fluxes during wetland forest re-treat along the lower Savannah River, Georgia (USA). Wetlands 32 (1), 73–81.

LaMontagne, M.G., Duran, R., Valiela, I., 2003. Nitrous oxide sources and sinks in coastalaquifers and coupled estuarine receiving waters. Sci. Total Environ. 309 (1), 139–149.

Larsen, L., Moseman, S., Santoro, A.E., Hopfensperger, K., Burgin, A., 2010. A complex-sys-tems approach to predicting effects of sea level rise and nitrogen loading on nitrogencycling in coastal wetland ecosystems. Eco-DAS VIII. 5, pp. 67–92.

Li, Y., Liu, M., Lu, M., Hou, L.J., Lin, X., 2010. Phragmites australis effects on N2O emission inthe Chongming eastern tidal flat (in Chinese). Acta Sci. Circumst. 30 (12), 2526–2534.

Liu, X., Han, J.G., Ma, Z.W., Wang, Q., Li, L.H., 2016. Effect of carbon source on dissimilatorynitrate reduction to ammonium in costal wetland sediments. J. Soil Sci. Plant Nutr. 16(2), 337–349.

Marton, J.M., Herbert, E.R., Craft, C.B., 2012. Effects of salinity on denitrification and green-house gas production from laboratory-incubated tidal forest soils. Wetlands 32 (2),347–357.

Moin, N.S., Nelson, K.A., Bush, A., Bush, A., Bernhard, A.E., 2009. Distribution and diversityof archaeal and bacterial ammonia oxidizers in salt marsh sediments. Appl. Environ.Microbiol. 75 (23), 7461–7468.

Morley, N.J., Richardson, D.J., Baggs, E.M., 2014. Substrate induced denitrification over orunder estimates shifts in soil N2/N2O ratios. PLoS One 9 (9), 1–6.

Moseman-Valtierra, S., Gonzalez, R., Kroeger, K.D., Tang, J., Chao, W.C., Crusius, John,Bratton, John, Green, A., Shelton, J., 2011. Short-term nitrogen additions can shift acoastal wetland from a sink to a source of N2O. Atmos. Environ. 45 (26), 4390–4397.

Mosier, A.C., Francis, C.A., 2008. Relative abundance and diversity of ammonia-oxidizingarchaea and bacteria in the San Francisco Bay estuary. Environ. Microbiol. 10 (11),3002–3016.

Musenze, R.S., Werner, U., Grinham, A., Udy, J., Yuan, Z., 2014. Methane and nitrous oxideemissions from a subtropical estuary (the Brisbane River estuary, Australia). Sci. TotalEnviron. 472, 719–729.

Musenze, R.S., Werner, U., Grinham, A., Udy, J., Yuan, Z., 2015. Methane and nitrous oxideemissions from a subtropical coastal embayment (Moreton Bay, Australia). J. Environ.Sci. 29, 82–96.

Musenze, R.S., Fan, L., Grinham, A., Werner, U., Gale, D., Udy, J., Yuan, Z., 2016. Methanedynamics in subtropical freshwater reservoirs and the mediating microbial commu-nities. Biogeochemistry 128 (1–2), 233–255.

Myhre, G., Shindell, D., Bréon, F.M., Collins, W., Fuglestvedt, J., Huang, J., Koch, D.,Lamarque, J.F., Lee, D., Mendoza, B., Nakajima, T., Robock, A., Stephens, G.,Takemura, T., Zhang, H., 2013. Anthropogenic and natural radiative forcing. In:Stocker, T., Qin, D., Plattner, G.K., Tignor, M., Allen, S., Boschung, J., Nauels, A., Xia,Y., Bex, V., Midgley, P. (Eds.), Climate Change. The Physical Science Basis. ContributionofWorking Group I to the Fifth Assessment Report of the Intergovernmental Panel onClimate Change. Cambridge University Press, Cambridge, United Kingdom and NewYork, NY, USA, p. 2013.

Osborne, R.I., Bernot, M.J., Findlay, S.E.G., 2015. Changes in nitrogen cycling processesalong a salinity gradient in tidal wetlands of the Hudson River, New York, USA. Wet-lands 35, 323–334.

Pina-Ochoa, E., Álvarez-Cobelas, M., 2006. Denitrification in aquatic environments, across-system analysis. Biogeochemistry 81 (1), 111–130.

Rahmstorf, S., 2007. A semi-empirical approach to projecting future sea-level rise. Science315 (5810), 368–370.

Please cite this article as: Wang, X., et al., Seasonal variations of nitrous oxbrackish tidal marshes of..., Sci Total Environ (2017), https://doi.org/10.10

Ravishankara, A.R., Daniel, J.S., Portmann, R.W., 2009. Nitrous oxide (N2O), the dominantozone-depleting substance emitted in the 21st century. Science 326 (5949), 123–125.

Santoro, A.E., 2010. Microbial nitrogen cycling at the saltwater-freshwater interface.Hydrogeol. J. 18 (1), 187–202.

Seitzinger, S.P., 1988. Denitrification in freshwater and coastal marine ecosystems, eco-logical and geochemical significance. Limnol. Oceanogr. 33 (4), 702–724.

Senga, Y., Mochida, K., Fukumori, R., Okamoto, N., Seike, Yasushi, 2006. N2O accumulationin estuarine and coastal sediments, the influence of H2S on dissimilatory nitrate re-duction. Estuar. Coast. Shelf Sci. 67 (1), 231–238.

Seo, D.C., Yu, K., Delaune, R.D., 2008. Influence of salinity level on sediment denitrificationin a Louisiana estuary receiving divertedMississippi River water. Arch. Agron. Soil Sci.54 (3), 249–257.

Silver, W.L., Herman, D.J., Firestone, M.K., 2001. Dissimilatory nitrate reduction to ammo-nium in upland tropical forest soils. Ecology 82 (9), 2410–2416.

Smith, C.J., DeLaune, R.D., Patrick, W.H., 1983. Nitrous oxide emission from Gulf Coastwetlands. Geochim. Cosmochim. Acta 47 (10), 1805–1814.

Strack, M., Waddington, J.M., 2008. Spatiotemporal variability in peatland subsurfacemethane dynamics. J. Geophys. Res. Biogeosci. 113:G02010. https://doi.org/10.1029/2007JG000472.

Sturm, K., Grinham, A., Werner, U., Yuan, Z., 2016. Sources and sinks of methane and ni-trous oxide in the subtropical Brisbane River estuary, South East Queensland, Austra-lia. Estuar. Coast. Shelf Sci. 168, 10–21.

Sun, W.G., Sun, Z.G., Gan, Z.T., Sun, W.L., Wang, W., 2014. Contribution of different pro-cesses in wetland soil N2O production in different restoration phrases of the YellowRiver estuary, China (in Chinese). Environ. Sci. 35 (8), 3110–3119.

Sun, W.G., Sun, Z.G., Mou, X.J., Sun, W.G., Hu, X.Y., 2017. Nitrous oxide emissions from in-tertidal zone of the Yellow River estuary in autumn and winter during 2011–2012.Estuar. Coasts 40, 145–159.

Tiedje, J.M., 1988. Ecology of denitrification and dissimilatory nitrate reduction to ammo-nium. Biol. Anaerob. Microorgan. 717, 179–244.

Tong, C., Huang, J.F., Hu, Z.Q., Jin, Y.F., 2013. Diurnal variations of carbon dioxide, methane,and nitrous oxide vertical fluxes in a subtropical estuarine marsh on neap and springtide days. Estuar. Coasts 36, 633–642.

Walter, S., Peeken, I., Lochte, K., Adrian, W., Bange, H.W., 2005. Nitrous oxide measure-ments during EIFEX, the European iron fertilization experiment in the subpolarSouth Atlantic Ocean. Geophys. Res. Lett. 32, 1–4.

Wang, D.Q., Chen, Z.L., Wang, J., Xu, S.Y., Yang, H.X., Chen, H., Yang, L.Y., Hu, L.Z., 2007.Summer-time denitrification and nitrous oxide exchange in the intertidal zone ofthe Yangtze Estuary. Estuar. Coast. Shelf Sci. 73, 43–53.

Wang, Z.Y., Xin, Y.Z., Gao, D.M., Li, F.M., Morgan, J., Xing, B.S., 2010. Microbial communitycharacteristics in a degraded wetland of the Yellow River Delta. Pedosphere 20 (4),466–478.

Wang, X.M., Ren, H.C., Tong, C., 2014. Effect of salinity on production and emission ofgreenhouse gases in estuarine tidal wetlands, a review (in Chinese). Wetland Sci.12 (6), 814–820.

Wang, W., Sardans, J., Lai, D.Y.F., Wang, C., Zeng, C., Tong, C., Liang, Y., Penuelas, J., 2015.Effects of steel slag application on greenhouse gas emissions and crop yield overmul-tiple growing seasons in a subtropical paddy field in China. Field Crop Res. 171,146–156.

Wang, H., Liao, G., D'Souza, M., Yu, X., Yang, J., Yang, X., Zheng, T., 2016. Temporal and spa-tial variations of greenhouse gas fluxes from a tidal mangrove wetland in SoutheastChina. Environ. Sci. Pollut. Res. Int. 23 (2), 1873–1885.

Wanninkhof, R., 1992. Relationship between wind speed and gas exchange over theocean. J. Geophys. Res. 97 (C5), 7373–7382.

Webster, F.A., Hopkins, D.W., 1996. Contributions from different microbial processes toN2O emission from soil under different moisture regimes. Biol. Fertil. Soils 22 (4),331–335.

Weiss, R.F., Price, B.A., 1980. Nitrous oxide solubility in water and seawater. Mar. Chem. 8(4), 347–359.

Welti, N., Hayes, M., Lockington, D., 2016. Seasonal nitrous oxide and methane emissionsacross a subtropical estuarine salinity gradient. Biogeochemistry https://doi.org/10.1007/s10533-016-0287-4.

Wigley, T.M.L., Raper, S.C.B., 1992. Implications for climate and sea level of revised IPCCemissions scenarios. Nature 357 (6376), 293–300.

Woodroffe, C.D., Murray-Wallace, C.V., 2012. Sea-level rise and coastal change, the past asa guide to the future. Quat. Sci. Rev. 54, 4–11.

Wrage, N., Velthof, G.L., van Beusichem, M.L., Oenema, O., 2001. Role of nitrifier denitrifi-cation in the production of nitrous oxide. Soil Biol. Biochem. 33 (12), 1723–1732.

Yang, J., Zhang, G.L., Zhao, Y.C., Li, P.P., 2011. Production and emission of nitrous oxidefrom the esturine tidal flat around Jiaozhou Bay and their influencing factors (in Chi-nese). Acta Sci. Circumst. 31 (12), 2723–2732.

Zheng, X.H., Wang, M.X., Wang, Y.S., Shen, C.X., Zhang, W., Gong, Y.B., 1997. Impacts oftemperature on N2O production and emission (in Chinese). Environ. Sci. 18 (5), 1–5.

Zhu, Y.G., Wang, X.H., Yang, X.R., Xu, H.J., Jia, Y., 2014. Key microbial processes in nitrousoxide emissions of agricultural soil and mitigation strategies (in Chinese). Environ.Sci. 35 (2), 792–800.

Zimmerman, A.R., Benner, R., 1994. Denitrification, nutrient regeneration and carbonmineralization in sediments of Galveston Bay, Texas, USA. Mar. Ecol. Prog. Ser. 114,275–288.

ide fluxes and soil denitrification rates in subtropical freshwater and16/j.scitotenv.2017.10.175