This article was downloaded by: [Dicle University]On: 06 November 2014, At: 08:03Publisher: Taylor & FrancisInforma Ltd Registered in England and Wales Registered Number: 1072954 Registered office: MortimerHouse, 37-41 Mortimer Street, London W1T 3JH, UK

Click for updates

International Journal of Biodiversity Science,Ecosystem Services & ManagementPublication details, including instructions for authors and subscription information:http://www.tandfonline.com/loi/tbsm21

Comparing soil carbon pools and carbon gas fluxes incoastal forested wetlands and flooded grasslands inVeracruz, MexicoMaria E. Hernandeza, Jose Luis Marín-Muñizb, Patricia Moreno-Casasolac & VioletaVázqueza

a Red de Manejo Biotecnológico de Recursos, Instituto de Ecología, A.C. CarreteraAntigua a Coatepec 351, El Haya, Xalapa, Veracruz, Mexicob Centro de Investigaciones Tropicales, Universidad Veracruzana, Casco de la Ex-Hacienda Lucas Martín, Privada de Araucarias S/N. Col. Periodistas, AP. 525, Xalapa,Veracruz, Mexicoc Red de Ecología Funcional, Instituto de Ecología, A.C. Carretera Antigua a Coatepec351, El Haya, Xalapa, Veracruz, MexicoPublished online: 18 Jun 2014.

To cite this article: Maria E. Hernandez, Jose Luis Marín-Muñiz, Patricia Moreno-Casasola & Violeta Vázquez(2014): Comparing soil carbon pools and carbon gas fluxes in coastal forested wetlands and flooded grasslandsin Veracruz, Mexico, International Journal of Biodiversity Science, Ecosystem Services & Management, DOI:10.1080/21513732.2014.925977

To link to this article: http://dx.doi.org/10.1080/21513732.2014.925977

PLEASE SCROLL DOWN FOR ARTICLE

Taylor & Francis makes every effort to ensure the accuracy of all the information (the “Content”) containedin the publications on our platform. However, Taylor & Francis, our agents, and our licensors make norepresentations or warranties whatsoever as to the accuracy, completeness, or suitability for any purpose ofthe Content. Any opinions and views expressed in this publication are the opinions and views of the authors,and are not the views of or endorsed by Taylor & Francis. The accuracy of the Content should not be reliedupon and should be independently verified with primary sources of information. Taylor and Francis shallnot be liable for any losses, actions, claims, proceedings, demands, costs, expenses, damages, and otherliabilities whatsoever or howsoever caused arising directly or indirectly in connection with, in relation to orarising out of the use of the Content.

This article may be used for research, teaching, and private study purposes. Any substantial or systematicreproduction, redistribution, reselling, loan, sub-licensing, systematic supply, or distribution in anyform to anyone is expressly forbidden. Terms & Conditions of access and use can be found at http://www.tandfonline.com/page/terms-and-conditions

Comparing soil carbon pools and carbon gas fluxes in coastal forested wetlands and floodedgrasslands in Veracruz, Mexico

Maria E. Hernandeza,*, Jose Luis Marín-Muñizb, Patricia Moreno-Casasolac and Violeta Vázqueza

aRed de Manejo Biotecnológico de Recursos, Instituto de Ecología, A.C. Carretera Antigua a Coatepec 351, El Haya, Xalapa, Veracruz,Mexico; bCentro de Investigaciones Tropicales, Universidad Veracruzana, Casco de la Ex-Hacienda Lucas Martín, Privada deAraucarias S/N. Col. Periodistas, AP. 525, Xalapa, Veracruz, Mexico; cRed de Ecología Funcional, Instituto de Ecología, A.C. CarreteraAntigua a Coatepec 351, El Haya, Xalapa, Veracruz, Mexico

Wetlands play an important role in carbon cycling. Perturbation of these ecosystems by human activities causes changes inthe soil carbon storage and carbon gaseous emissions. These changes might have important repercussions for globalwarming. The aim of this study was to investigate whether the conversion of freshwater forested wetlands (FW) to floodedgrasslands (FGL) has affected soil carbon cycling. Soil carbon pools and soil organic carbon (SOC) fractions (water-solublecarbon (WSC), hot-water-soluble carbon (HWSC), and HCl/HF soluble carbon (HCl/HF-SC)) were compared between FWand FGL. Additionally, the seasonal dynamic of methane (CH4) and carbon dioxide (CO2) fluxes were monitored in bothecosystems located in the coastal plain of Veracruz State Mexico. In FW, soil organic matter (SOM) concentrations weresignificantly (P ≤ 0.05) higher than FGL. Soil bulk density (BD) was slightly higher in FGL than FW but it wasnot significantly different (P ≥ 0.05). The average of WSC and HWSC in FW were not significantly (P ≤ 0.05) different.Total carbon pools (44 cm deep) were not significantly different (P = 0.735). During the dry season, CO2 fluxes(26.38 ± 4.45 g m−2 d−1) in FGL were significantly higher (P = 0.023) than in FW (14.36 ± 5.77 g m−2 d−1). During therainy and windy seasons, both CH4 and CO2 fluxes were significantly higher (P = 0.000 and P = 0.001) in FGL comparedwith FW. It was concluded that converting FW to FGL causes loss of SOC and increases carbon gaseous fluxes.

Keywords: carbon cycle; global warming; land-use change; soil organic carbon; tropical freshwater wetlands; ecosystemservices

1. Introduction

Wetlands are the interface between terrestrial and aquaticcomponents of the landscape (Mitsch & Gosselink 2007).They are widely recognized for providing several ecosystemservices such as flood control, aquifer recharge, and nutrientremoval (Hansson et al. 2005). However, their contributionto the global cycling of atmospheric gases and their impor-tant role as carbon sinks is less recognized. Wetlands, despiteoccupying relatively small areas of the earth’s surface (2–6%), contain a large proportion of the world’s carbon storedin terrestrial soil reservoirs (Whiting & Chanton 2001; Mitraet al. 2005; Lal 2008). Soil organic carbon (SOC) pool iscomplex; based on resistance to mineralization, it has beendivided into labile, intermediate, and recalcitrant organiccarbon pools (Cheng et al. 2007). Soil labile and intermediateC pools have a mean residence time of years to severaldecades while recalcitrant C pools have a mean residencetime of hundreds to thousands years (Zou et al. 2005; Chenget al. 2007; Silveira et al. 2008). Labile and intermediatefractions of organic carbon can respond rapidly to environ-mental change; therefore, they are more sensitive indicatorsof the effects of land use than total SOC (von Lützow et al.2002; Zhang et al. 2007; He et al. 2008).

In wetlands ecosystems, flooding conditions not onlyallow accumulating significant amounts of carbon, but also

promote the production and release of methane (CH4), apowerful greenhouse gas (GHG). Wetlands are a majorsource of CH4 in the atmosphere (Whalen 2005) contributing23–40% of the annual terrestrial CH4 emissions and compris-ing 77–83% of natural sources (IPCC 2001). In addition toCH4, carbon dioxide (CO2) is produced in wetlands soilsunder both aerobic and anaerobic conditions (Smith et al.2003; Coles & Yavitt 2004; Elberling et al. 2011). CH4

emissions to the atmosphere are an environmental concernbecause global warming potential (GWP) for CH4 is 25 timesthan GWP for CO2 (Solomon et al. 2007). Therefore, smallincrease of CH4 concentration in the atmosphere might havean important impact on global warming.

Human activities can alter the carbon stocks in wet-lands and the exchange of GHG with the atmosphere(Roulet 2000). For example, it has been reported thatlivestock grazing, significantly reduced the above-groundbiomass, net primary productivity, and enhanced CH4

emissions in wetlands on the Qinghai-Tibetan plateau inChina (Hirota et al. 2005). According to Solomon et al.(2007), global increases in CO2 concentration are dueprimarily to land-use change and fossil-fuel consumption.Based on measured CO2 fluxes using satellite observationsand emission inventories in China, Wang et al. (2011)reported almost five times more CO2 emissions in slope

*Corresponding author. Email: elizabeth.hernadez@inecol.mxNotes: FW = forested wetlands, FGL = flooded grasslands, WSC = water-soluble carbon, HWSC = Hot-water soluble carbon,HCl/HF-SC = HCl/HF soluble carbon, BD = bulk density, SOC = soil organic carbon, SOM = soil organic matter.

International Journal of Biodiversity Science, Ecosystem Services & Management, 2014http://dx.doi.org/10.1080/21513732.2014.925977

© 2014 Taylor & Francis

Dow

nloa

ded

by [

Dic

le U

nive

rsity

] at

08:

03 0

6 N

ovem

ber

2014

grasslands compared with swamp soils, considering simi-lar areas. However, despite the potential importance, fewstudies are available to assess the effects of livestockgrazing on the GHGs emissions in tropical wetlandecosystems.

In Mexico, wetlands are located mainly on the coast(Contreras-Espinosa & Warner 2004), and these areas areamong the most transformed ecosystems in the country(Moreno-Casasola 2008). Coastal wetlands in the Mexicanstate of Veracruz are threatened mainly by cattle ranching,petrochemical activities, and urbanization. Because of this,most freshwater wetlands show significant changes in theirecology, such as invasion of exotic species, siltation, pol-lution, and changes in their hydrology (López-Rosas et al.2006; Moreno-Casasola et al. 2009).

In order to assess the impact of human activity on thecarbon sequestration service that wetlands provide; theobjective of this study was to compare soil organic matter(SOM), soil carbon pools, and the seasonal carbon gaseousfluxes (CO2 and CH4) in both coastal freshwater swampsand areas that have been converted to flooded grasslands(FGL). We hypothesize that carbon pools in flooded grass-lands will be lower than in forested wetlands (FW) due tothe decrease of carbon inputs and higher mineralizationrates. The opposite will occur with carbon gaseous fluxes,that is, higher fluxes in FGL due to higher carbonmineralization.

2. Materials and methods

2.1. Study site

The study was carried out in two freshwater FW and twoadjacent flooded grasslands, located on the coastal plain ofthe Gulf of Mexico in the state of Veracruz. The study sites

located from north to south were Estero Dulce(20º17ʹ53ʺN, 96º52ʹ19ʺW) and Boquilla de Oro(19º49ʹ47ʺN, 96º26ʹ59ʺW) (Figure 1). Flooded grasslandswere established in FW 15–20 years ago. Natural wetlandshave been transformed to support cattle ranching as eco-nomic activity. The transformation included cutting nativewetland trees to allow the growth of native and introducedflood tolerant grasses. Drainage of these areas has not beenperformed. However, the introduction of exotic grassescauses that flooded grasslands experience shorter hydro-periods than natural wetlands (López-Rosas et al. 2006).The studied FW are fenced to exclude grazing, whileflooded grasslands experience from moderate (1.8 animalsper ha) to heavy (3 animals per ha) grazing (Girma et al.2007). These areas are grazed from March to early August,and afterward animals are moved to uplands because ofthe rainy season. Flooded grasslands are not fertilized withchemicals; and tillage is not performed in these sites. Adetailed description of the study sites is shown in Table 1.

The climate of the coastal plain of the Gulf of Mexicohas three seasons: rainy season (July to October), windyseason (November to February) that has cold fronts withstrong winds and rain, and the dry season (March to June).The annual precipitation mean fluctuates between 1200and 1650 mm. The mean annual temperature variesbetween 17°C and 37°C. A detailed description of thestudy sites is shown in Table 1.

2.2. Soil sampling

In each type of wetland (FW or FGL), three randomsampling plots (1 m2) were established. In these plots,four soil cores (0.48 m deep × 0.05 m diameter) weretaken using a Russian peat borer. This borer has thin

97°0'0''W 96°0'0''W 95°0'0''W

20°0

'0''W

20°0

'0''W

95°0'0''W96°0'0''W

0 10 20 40 60 80Km

97°0'0''W

Figure 1. Location of the study sites in the coastal plain of Veracruz, Mexico.

2 M.E. Hernandez et al.

Dow

nloa

ded

by [

Dic

le U

nive

rsity

] at

08:

03 0

6 N

ovem

ber

2014

sharp-edged walls, providing a core without compaction,distortion, or disturbance. Each core was sectioned offwith a blade at intervals of 4 cm. One of the four coresin each sampling plot was used for analysis of bulk density(BD). The soil layers were placed in aluminum pans hav-ing a predetermined dry weight, and transported to thelaboratory where they were stored below 4ºC until theywere dried in an oven.

Composite samples were made with three wet soil sam-ples taken from the same depth in each sampling plot. Eachof the mixed wet composted samples were packed in con-tainers and stored below 4ºC until they were dried at roomtemperature and analyzed for carbon content.

2.3. Soil analysis

In the laboratory, wet soil composite samples were mixedagain to reduce soil heterogeneity and visible residues ofvegetation were removed. Composite samples were driedat room temperature, pulverized, and sieved (2 mm). Forquantifying the organic matter, approximately 2 g of driedsoil samples were pretreated with 10 M HCl to avoidpossible carbonate interferences (Hernandez & Mitsch2007). After this, SOM was quantified by loss on ignitionat 450ºC for 4 hours (Craft et al. 1988; Bernal & Mitsch2008).

BD was obtained by drying a known volume of sedi-ment at 105ºC (19.64 cm−3); then, it was weighed until aconstant weight was reached. The values obtained wereused in the formula BD (g cm−3) = Mass/volume.

For the purpose of carbon pool calculations, theorganic carbon percentage was calculated as a portion oforganic matter, using Van Bemmelen’s factor (0.58) whichhas been used for several wetland soils including thesetropical wetlands (Wang et al. 2003; Hernandez & Mitsch2007; Marín-Muñiz et al. 2014). The carbon pool wascalculated in kg C m−2, according to the following equa-tion (Moreno et al. 2002; Cerón-Bretón et al. 2011):

Kg C m�2 ¼ soil dry weight½ � � OC½ �

where: soil dry weight (kg m−2) = [sampled soil depth] *[bulk density], and OC = organic carbon content.

Total carbon storage to a 44 cm depth was calculatedby adding the carbon stored in each one of the soil layers(Bernal & Mitsch 2012).

2.4. Soluble organic carbon fractions

Extractions of water-soluble carbon (WSC), hot-water-soluble carbon (HWSC), and HCl/HF soluble carbon(HCl/HF-SC) were carried out according to Hernandezand Mitsch (2007). Soluble organic carbon concentrationsin each of the extracts were analyzed in a Total OrganicCarbon analyzer (Torch, Teledyne Tekmar).

2.5. Gas measurements and flux calculations

Fluxes of CH4 were measured in situ once every 2 monthsstarting in August 2010 until February 2012, and the CO2

from February 2011 to February 2012, using the closedchamber technique (Altor & Mitsch 2006; Hernandez &Mitsch 2006; Nahlik & Mitsch 2010). The closed chamberconsisted of two parts: a base and a removable cap, eachmade of polyvinyl chloride (PVC) pipe (15 cm diameter).The bases were permanently installed in the swamps inFebruary 2010 and in June 2010 in the flooded grasslands(four chambers in each type of wetland at the two sites,n = 8 for each type of wetland). The bases were 30 cmhigh and inserted approximately 5 cm into the wetlandsoils; the base had an open bottom and a collar, 5 cm fromthe top. The removable cap includes a gray butyl samplingport and an alcohol-type thermometer in the top. Everytime gas fluxes were measured, the cover was put on thebase collar, and water was added to ensure a gas-tight sealbetween the base and the cap. Chambers were closed, andevery 5 minutes internal gas samples were taken for thenext 45 minutes and the internal temperature registered.Gas samples (25 ml) were taken using 60-ml propylenesyringes (TERUMO) having a one-way stopcock (Lieur).Gas samples were injected through rubber septa into pre-

Table 1. Characteristics of the studied wetlands in the coastal plain of Veracruz, Mexico.

Study site Location Ecosystem type Dominant plant speciesAnimalsper ha

No of years oftransformation

to FGL Soil typea

Estero Dulce(ED)

20º17ʹ53ʺN 96º52ʹ19ʺW Forested wetland Pachira aquatica Aubl 0 0 HisticGleysol

Floodedgrassland

Cynodon plectostachyusCladium jamaicense

3 More than 15

Boquilla deOro (BO)

19º49ʹ47ʺN, 96º26ʹ59ʺW Forested wetland Ficus insipida andPleuranthodendronlindenii (Turcz.) Sleumer

0 0 SapricHistosol

Floodedgrassland

Cynodon plectostachyusCladium jamaicense

2 More than 20

Note: aData from previous studies in the area (Campos et al. 2011; Infante et al. 2012).

International Journal of Biodiversity Science, Ecosystem Services & Management 3

Dow

nloa

ded

by [

Dic

le U

nive

rsity

] at

08:

03 0

6 N

ovem

ber

2014

evacuated 20-ml glass vials. Septa were boiled before usefor 30 minutes to eliminate potential gas leaking. Allsamples were taken between 10:00 h and 16:00 h (localtime) and were analyzed within 72 h after collection.

Gas concentrations were analyzed on a Perkin ElmerClarus 500 gas chromatograph equipped with a flame ioniza-tion detector (FID) for CH4 and also equipped with a metha-nizer to detect levels of CO2. For sample separation, astainless steel column packed with Poropak Q (80/100mesh), 6 ft in length, and 2-mm ID was used. The tempera-tures for oven, injector, and detector were at 40, 95, and 200°C for CH4, and 60, 80, and 350°C for CO2, respectively.Nitrogen (7 ml min−1) was used as a carrier gas. CH4 andCO2 were quantified separately. Matheson gas standardsbalanced with N2 were used to perform calibration curves.All the individual analyzed gas values (ppm CH4 and CO2)were corrected using the ideal gases law (pv = nRT) accord-ing to the formula (Duan et al. 2009; Nahlik &Mitsch 2010):

m ¼ c� P �Mð Þ= T � Rð Þ

where: m is the gas concentration (g m−3), c the gasconcentration by volume (ppmv (cm3 m−3)), P the atmo-spheric pressure (assume 1 atm), M the molecular weightof gas (g mol−1), R the Universal Gas Constant(82.0576 (atm.cm3)/(mol-K)), and T the air temperature(K) inside the chamber at the time of each sample.

The normalized gas concentrations were used to calcu-late gas flux rates (Hernandez & Mitsch 2006) accordingto the following equation:

Fc ¼ dc=dtð Þ � V=Að Þð Þ � 1440

where: Fc = is the flux rate (mg m−2 d−1), (dc/dt) = change ingas concentration over the enclosure period, expressed as(mg m−3 min−1), V the chamber volume (m3), and A the basechamber soil-surface area (m2), 1440 = minutes by day.

For each chamber measurement, gas sample concen-tration values were plotted versus sample time. MicrosoftExcelTM was used to calculate linear regressions on eachflux rate. Results were included only if R2 was greater than0.85 (Altor & Mitsch 2006).

2.6. Conversion to CO2-equivalents

A GWP factor of 25 for CH4 (Solomon et al. 2007) wasused to convert CH4 emissions to CO2-equivalents forcomparing their contributions to the global radiativeimpact.

2.7. Physical and chemical analysis

2.7.1. Water level

When surface water was present, the water level wasrecorded using a measuring stick. When no surface waterwas present, a sensor connected to a multimeter (Steren)was used to detect the water level in four monitoring

wells, located in each type of wetland at the three studysites. Monitoring wells were made from PVC pipe (13 mmID), 3 m in length (inserted 1.5 m in the soil), installed ineach studied wetland (Infante et al. 2012).

2.7.2. Redox potential

Soil redox potential (Eh) was measured within a 30 cmdiameter around each chamber, at a soil depth between 0and 5 cm using a platinum rode and one calomel referenceelectrode (Corning 476,340), both connected to a digitalmultimeter. Platinum electrodes were calibrated in situbefore every monitoring with quinhydrone (Aldrich)50 mgl−1 in a pH 4.0 buffer solution (Bohn 1971).

2.8. Statistical analysis

All statistical analyses were performed with SPSS version18 for Windows. A Kolmogorov–Smirnov test was usedto check normality. Physical–chemical variables andcarbon in soil data fit normal distributions. One-wayanalysis of variance (ANOVA) was used to find outwhether the type of ecosystem had an effect on theaverage of soil carbon concentration, BD, WSC, HWSC,and HCl/HF-SC. These parameters were compared at thesame depth between FW and FGL using a t-test for pairedsamples. To detect differences in carbon pools betweentwo types of ecosystems, a t-test was used. Two-wayANOVA with Tukey comparison was used to determinewhether climatic season and ecosystem types had an effectin Eh, water levels, and soil temperature. Data for GHGsfailed to meet criteria for normal distribution (P < 0.001);therefore, non-parametric statistical tests were used suchas Kruskal-Wallis and Mann-Whitney to compare CH4

and CO2 emissions between the type of ecosystem andclimatic season. A non-parametric t-test for paired sampleswas used to compare the emissions of CH4 or CO2

between FW and FGL during the same months of sam-pling. A p-value = 0.05 was used to reveal the statisticalsignificance.

3. Results

3.1. Soil organic matter concentration, soil bulkdensity, and carbon pools

The SOM concentration decreased in the studied FWsoils according to the depth from 382 to 300 g kg−1

(Figure 2), while in the FGL, SOC decreased from 172to 97 g kg−1. When SOM concentration was comparedbetween FW and FGL at the same depth, they were sig-nificantly higher (P ≤ 0.05) in FW, for all depths. AverageSOM in FW (284.25 ± 15.2 g kg−1) was also significantlyhigher (P = 0.001) than average SOM in FGL(134.41 ± 6.33 g kg−1).

Soil BD in FGL varied from 0.44 to 0.77 g cm−3,increasing with depth, while BD in the FW varied from0.41 to 0.57 g cm−3; and BD did not increase with depth.

4 M.E. Hernandez et al.

Dow

nloa

ded

by [

Dic

le U

nive

rsity

] at

08:

03 0

6 N

ovem

ber

2014

Although we observed a trend of higher BD in FGL,when it was compared to FW, no significant differenceswere found (P ≥ 0.05) neither at any depth, nor in theaverage between FGL (0.63 ± 0.02 g cm−3) and FW(0.49 ± 0.02 g cm−3).

When the total carbon pools were calculated to 44 cm,the results were 30.63 ± 5.23 kg C m−2 for FGL and28.14 ± 4.87 kg C m−2 for FW, showing no significantdifference (P = 0.735).

3.2. Water-soluble C (WSC), hot-water-soluble C(HWSC), and HCl/HF soluble C

In FGL soils, WSC decreased with depth from 0.70 to0.27 g C kg −1, while in FW soils, WSC increased after17 cm from 0.50 to 0.90 g C kg −1. In the top layer, WSCwas slightly higher in FGL than FW but not significantlydifferent (P = 0.094). Average WSC in the whole soilprofile of FGL (0.74 ± 0.08 g C kg −1) was also notsignificantly different (P = 0.094) from FW(0.54 ± 0.09 g C kg −1) (Figure 3).

HWSC decreased with depth in both ecosystems; FGLshowed higher HWSC than FW in the whole profileexcept in the deepest layer. The average of HWSC in thewhole profile of FGL was 1.19 ± 0.19 g C kg −1, while inFW it was 1.01 ± 0.15 g C kg −1 being not significantlydifferent (P = 0.309).

Organic carbon extracted by HCl/HF decreased sig-nificantly with depth in both types of wetlands(P < 0.05). HCl/HF-SC in FW was higher than inFGL in the entire profile. Average of HCl/HF-SC inthe entire profile of FGL (0.29 ± 0.05 g C kg −1)was significantly lower (P = 0.006) than in FW(0.41 ± 0.07 g C kg −1).

3.3. Water level, redox potential, and soil temperaturedynamics

Water level in the studied wetlands ranged from −70 to10 cm (Figures 4a–b) in FW and from −45 to 10 cm inFGL, without significant differences (Figure 4a;P = 0.998). When water level values were averaged pereach season, significant differences were observed(P = 0.021) with higher values for the rainy (4.86 ± 2.30and 1.18 ± 3.89 cm in FGL and FW, respectively) and thewindy seasons (−4.30 ± 3.99 and 3.53 ± 1.79 cm in FGLand FW, respectively), compared with water levelsobserved during the dry season (−33.75 ± 10.78 cm inFGL and −37.99 ± 24.24 in FW).

Eh values in the soil oscillated from −37 to 350 mV inFW and from 0 to 337 mV in FGL (Figures 4c–d), and nosignificant differences were observed (Figure 4c;P = 0.613). When Eh values were averaged by season, Ehaverage in FW decreased from dry (216.26 ± 136.83 mV) to

0

(a) (b)

4

X– = 134.4 ± 6.3

g kg–1

X– = 284.3 ± 15.2

g kg–1X– = 0.49 ± 0.02

g cm–3

X– = 0.63 ± 0.02

g cm–3

FGLFW

8

12

16

20

24

Dep

th (

cm)

28

32

36

40

44

100 200 300 400 0 0.2 0.4

Bulk density(g cm–3)

Organic matter(g kg–1)

0.6 0.8 1

X– = 134.4 ± 6.3

g kg–1

X–

= 284.3 ± 15.2g kg–1

X– = 0.49 ± 0.02

g cm–3

X– = 0.63 ± 0.02

g cm–3

FGLFW

Figure 2. Profile of soil organic matter (a) and bulk density (b) in the FW (gray lines) and FGL (black lines) soils. Each point in thegraph is the mean of six composite cores at 4 cm of depth. Horizontal bars represent standard errors.

International Journal of Biodiversity Science, Ecosystem Services & Management 5

Dow

nloa

ded

by [

Dic

le U

nive

rsity

] at

08:

03 0

6 N

ovem

ber

2014

rainy (63.07 ± 42.16 mV) and windy–rainy seasons(170.98 ± 52.13 mV), while in the FGL, Eh values did notdecrease (173.03 ± 164.09, 185.12 ± 6.17 and

209.22 ± 24.45 mV, respectively). In both types of wetlandsEh values were not significantly different among the sea-sons (P = 0.695).

0.0

7

17

27

X–

= 0.54 ± 0.09g C kg–1

X–

= 0.74 ± 0.08g C kg–1

X–

= 1.19 ± 0.19g C kg–1

X–

= 1.01 ± 0.15g C kg–1

g C kg–1

X–

= 0.29 ± 0.05g C kg–1

X–

= 0.41 ± 0.07g C kg–1

FGL

FW

Dep

th (

cm)

37

47

0.5 1.0 1.50.0 1.0 1.5 2.0

b)HWSCa)WSC

2.50.0 0.2 0.4 0.6 0.8

HCL/HF-SC c)

Figure 3. Carbon fractions in the FW wetland (gray lines) and FGL (in black lines) soils. Horizontal bars represent standard errors.

Figure 4. Water levels (a–b), redox potential (c–d), and soil temperature (e–f), in FGL (○) and FW (●) measured bimonthly (left) andaveraged by climatic season (right); white bars are flooded grasslands, gray bars are FW. Vertical lines on bars and circles values representstandard error, and different letters indicate significant difference.

6 M.E. Hernandez et al.

Dow

nloa

ded

by [

Dic

le U

nive

rsity

] at

08:

03 0

6 N

ovem

ber

2014

Soil temperatures in the freshwater wetlandsranged from 19 to 37°C in both types of wetlands(Figures 4e–f). Soil temperatures in FGL were 1–2°Chigher than in FW during the study period with few excep-tions. However, no significant differences were foundbetween the two types of wetlands (P = 0.271). Seasonshad a significant effect (P = 0.020) on soil temperature withlower temperatures during the windy season (19–21ºC) ascompared to rainy (27–28ºC) and dry seasons (28–32ºC).

3.4. Methane and carbon dioxide emissions

Both CH4 and CO2 fluxes were significantly influenced bythe season (P = 0.0001, P = 0.0001) with higher CH4

emissions during the rainy (August–October) and windy(December–February) seasons compared with the dry sea-son (April–June) (Figure 5). For CO2, the oppositeoccurred; high fluxes were observed during the monthsof dry season and low fluxes during rainy and windyseasons. Additionally, both gas emissions were signifi-cantly affected by the type of ecosystem. During the dryseason, average emissions of CO2 (26.38 ± 4.45 g m−2 d−1) were significantly higher (P = 0.023) in FGL comparedwith FW (14.36 ± 5.77 g m−2 d−1), while CH4 emissionswere low and similar in both types of ecosystems

(150.14 ± 75.22 and 145.68 ± 30.47 mg m−2 d−1 in FWand FGL, respectively; P = 0.224). During the rainy sea-son, significantly higher CH4 (P = 0.000) and CO2 emis-sions (P = 0.001) were found in FGL(4349.03 ± 853.46mg m−2 d−1 and 11.82 ± 1.24 g m−2

d−1, respectively) than in FW (869.01 ± 314.27mg m−2 d−1

and 4.59 ± 1.87 g m−2 d−1, respectively). Also in thewindy season, significantly higher CH4 (P = 0.001) andCO2 (P = 0.014) emissions were found in FGL(3912.01 ± 1378.30mg m−2 d−1 and 10.3 ± 8.07 g m−2

d−1, respectively) than in FW (481.66 ± 324.92mg m−2 d−1

and 2.98 ± 2.77 g m−2 d−1, respectively).

3.5. Global warming potential

We converted seasonal average emission of CH4 intoCO2-equivalents to compare its cumulative contributionsto global radiative balance (Figure 6). During the dryseason the main component of GHG fluxes was CO2

flux for both types of ecosystems. For rainy and windyseasons the main component of GHG fluxes was CH4 forboth types of ecosystems. FGL had statistically higher(P ≤ 0.05) radiative balance in all rainy and windy sea-sons (120 and 108 g m−2 d−1, respectively) than FW (26and 15 g m−2 d−1, respectively). On the other hand, during

Figure 6. Total emissions of methane and carbon dioxide expressed as CO2-equivalents according to GWP (CH4:25 and CO2:1;Solomon et al. 2007).

Figure 5. Methane (gray line) and carbon dioxide (black line) emissions in FGL and FW soils. Vertical lines represent standard errors.

International Journal of Biodiversity Science, Ecosystem Services & Management 7

Dow

nloa

ded

by [

Dic

le U

nive

rsity

] at

08:

03 0

6 N

ovem

ber

2014

dry season the radiative balance was not significantly(0.985) different between FGL (30 g m−2 d−1) and FW(18 g m−2 d−1).

4. Discussion

In the areas converted to FGL, SOC was only 47% of theobserved in FW. This is probably due to the decrease inthe carbon inputs to the soils and changes in hydrology.Litterfall in these tropical FW has been described as highas 9–15 ton ha per year−1 (Infante et al. 2012) and a largepart of the carbon remains in the flooded soils. However,parts of these FW were transformed to FGL at least15 years ago (Table 1). The transformation included clear-ing trees that allowed the growth of native and exoticflood-tolerant grasses to support cattle grazing.Channelization to drain FGL has not been performed inthese sites; however, changes in vegetation cover haveinduced shorter hydroperiods in FGL than in FW(Figure 4). Even though water levels were not significantlydifferent, FW were flooded during windy–rainy seasonwhile the FGL experienced some dry periods. Shorterhydroperiods might stimulate high C mineralization as itwas confirmed by observed high carbon fluxes in thisstudy. The loss of organic carbon in degraded wetlandsdue to changing land use has been described in othertropical wetlands. Sigua et al. (2009) found that naturalwetlands in South Florida had 180 g C kg−1 while alteredpastures (wetlands converted to pastures for 45 years) hadonly 5.4 g C kg−1 and after 6 years of wetland restorationSOC increased to 10.7 g C kg−1. On the other hand, intemperate wetlands, Shang et al. (2013) described the term‘grasslandification’ as the process where wetlands aredrained and converted to pasture with dominant plants ofgrasses. They described that grasslandification for 50 yearsin Chinese alpine wetlands had reduced vegetation qualityand increased the degree of drought and reduced the C, N,and P content of soils. Their observations are similar to thefindings in transformed FW to FGL for at least 15 years inVeracruz, Mexico. There are few studies of soil carbon inMexican wetlands (Campos et al. 2011; Marín-Muñizet al. 2014), and this study is the first report comparingSOC in disturbed wetlands. In other Mexican ecosystemssuch as the upland tropical forest ‘La selva Lacandona’,land-use change to pasture has decreased soil carbon poolsby approximately 50% (De Jong et al. 2000). In theBrazilian Amazonia, conversion of forest to pastures overseveral decades also caused a decrease of SOC (Fearnside& Barbosa 1998).

Despite of low organic carbon content in FGL, carbonpools were similar to FW. This happens because carbonpools were calculated using SOC content and BD, and thelatter were higher in FGL than the observed in FW. HigherBD in FGL might be caused by compaction due to cattlehoof action and shorter hydroperiods (Howe et al. 2009;Teuber et al. 2013). Similar results were obtained in Zoigealpine wetlands in China; degraded wetlands (floodedmeadows) had higher carbon pools than pristine FW

despite the latter had higher carbon concentrations butlower BD (Huo et al. 2013).

The simplest methods to measure available C sub-strates or labile carbon in both agricultural and wetlandsoils are WSC and HWSC. Land-use changes cause soildegradation, and sometimes these carbon fractions aremore sensitive than total organic carbon to such degrada-tion (Ghani et al. 2003; Dodla et al. 2012; Uchida et al.2012). We measured WSC and HWSC to investigatewhether the transformation of FW to FGL has affectedthe carbon cycling. However, no clear differences wereobserved. A trend of higher WSC in the top layer of FGLwas found but in deeper layers FW had higher concentra-tions of WSC. The high concentration of WSC in deeperlayers in wetland soil is due to leaching of WSC from thetop layer due to flooding conditions (Dodla et al. 2012). Inthis study, FGL showed shorter hydroperiods which mighthave limited the leaching of WSC. Although not statisti-cally different, HWSC also showed a trend of higherconcentrations in the whole profile of FGL than in FW.HWSC consists of a labile pool of SOM which includesmicrobial biomass as well as soluble soil carbohydratesand amines (Ghani et al. 2003). The fact that this type ofcarbon was higher in FGL and might also due to shorterhydroperiods that enhance less reduced conditions andhigher activity of aerobic microorganisms that hydrolyzeSOM releasing HWSC. We found that HWSC were threeto four times higher than WSC, and this occurs becausehot water dissolves more complex carbon compounds suchas microbial biomass C, root exudates, amino acids, and Cbound to soil enzymes. The results of this study are similarto the values found in coastal wetland soils of theMississippi River deltaic plain where HWSC was 4–13higher than WSC (Dodla et al. 2012). Also, Ghani et al.(2003) found in uplands soils that WSC constituted onlyapproximately 3–6% of HWSC and this type of carbon hasbeen correlated positively with soil respiration (Uchidaet al. 2012). In this study, the results showed a highercontent of HWSC in FGL than in FW, and the formerhas the higher carbon gaseous fluxes. On the other hand,the carbon fraction – HCl/HF-SC was significantly higherin FW than in FGL. This fraction represents carbon clo-sely associated with soil minerals (Al and Fe) and it isconsidered less available for microorganisms than WSCand HWSC (Stevenson 1982; Nguyen 2000). Other stu-dies in upland soils have indicated that Al- and Fe-boundorganic matter fractions were subjected to depletion duringthe harvesting and pasturing (Murata et al. 1995).

We found a strong seasonal influence on carbon gas-eous fluxes in both FW and FGL. When water tablesdropped during the dry season both types of wetlandsshowed low CH4 emissions and high CO2 emissions. Incontrast, when soils were flooded (rainy and windy sea-sons), higher CH4 emissions and lower CO2 emissionswere observed. This finding agrees with several studiesthat have described that CH4 emissions are favored whensoils are flooded (Altor & Mitsch 2006; Nahlik & Mitsch2010; Morse et al. 2012). However, despite both types of

8 M.E. Hernandez et al.

Dow

nloa

ded

by [

Dic

le U

nive

rsity

] at

08:

03 0

6 N

ovem

ber

2014

wetlands showed the same seasonal trend in carbon gas-eous fluxes, the magnitude of CH4 and CO2 fluxes in FGLwas higher than in FW. This agrees with results found indisturbed wetlands. Hirota et al. (2005) found that live-stock grazing stimulated CH4 emissions from alpine wet-lands in Tibet, compared with wetlands without grazing.Also, Oates et al. (2008) observed greater CH4 emissionsunder grazing conditions in spring-fed wetlands of aCalifornia oak savanna. In this study, the explanation forhigher CH4 emissions in FGL compared with FW mightbe due to several factors, including the physical soil dis-turbances by hoof action of cattle, changes in vegetationcover, different hydroperiods, and changes in the soil’schemistry due to deposition of cattle excreta on soilsduring grazing.

Cattle disturb soil porosity and break up the stratifi-cation of surface and sub-surface water, which contain O2

and methanotrophic bacteria. It has been described thatlivestock grazing and agricultural practices may have aneffect on the soil’s ability to consume CH4 by altering thedistribution of pore space, thereby reducing CH4 diffu-sion rates through the soil profile, and slowing transportto sites of CH4 oxidizing bacteria (Boeckx & Cleemput1997). Compaction leads to a reduction of aerobic micro-sites and consequently the decrease of CH4 oxidation byoxidizing bacteria (Sitaula et al. 2000). Our study foundgreater compaction in FGL than in FW and is potentiallya major influence on reduction of CH4 consumption.Besides these physical factors, the change from nativewetland trees to grasses in FGL might also havedecreased CH4 oxidation, because wetland plants supplyoxygen to the rhizosphere, which enhances areas ofpotential CH4 oxidation in the soil (Brix et al. 1996;Frenzel & Rudolph 1998).

Hydrology is one of the factors controlling Eh whichinfluences biogeochemical process in wetlands soils(Mitsch & Gosselink 2007). The sediment Ehs in thestudied sites were moderately reducing (−100 to250 mV) (Bohn 1971). Methanogenesis is such an obligateanaerobic process that it would not be expected to occur insediments until the Eh is at least −150 mV (Wang et al.1993; Kludze & DeLaune 1994). However, authors suchas Huang et al. (2005) and Wang et al. (1993) also havefound that methanogenic activities are still active at valuesclose to −100 mV. Chapelle et al. (1996) described thatalthough Eh measurements are easy to do in the field, theydo not always indicate with accuracy the anoxic biogeo-chemical process in the soils; and this is one possibleexplanation to the results in this study.

Paradoxically, higher CH4 emissions were found inFGL, which have shorter hydroperiods and although notstatistically a distinguishable high Eh compared with FW.This might be due to more wet and dry cycles that FGLexperience in comparison with FW, especially during rainyand windy seasons (Figure 4a). Badiou et al. (2011)described that CH4 emissions in the restored wetlands ofthe Canadian prairie pothole region increased dramaticallyjust as the wetland basin was becoming dry. The same

trend was observed by Pennock et al. (2010) in an ephem-eral wetland in Saskatchewan, Canada. Badiou et al.(2011) described that the transition period causes therelease of a massive pulse of CH4 due to the fact that thewetland sediment is still saturated and anoxic favoringmethanogenesis. Additionally, the sediment surfacewould warm dramatically as water levels decrease, therebyincreasing rates of methanogenesis which are known toincrease with soil temperature (Bartlett & Harriss 1993).Lastly, the decrease in water column depth above thesediment–water interface would facilitate the transfer ofCH4 from the sediments to the atmosphere while reducingthe potential for CH4 consumption within the water col-umn. The more frequent wet and dry cycles in FGL mightalso be responsible for higher CO2 emissions than FWduring the rainy and windy seasons. Wilson et al. (2011)found that carbon mineralization and therefore CO2 emis-sions increased significantly after flooding occurred inriparian wetland soils.

Soil fertilization caused by cattle excreta deposition onFGL soil is another factor that might explain the higherCO2 and CH4 emissions in these sites. Studies in uplandssoils have shown that manure addition to soils increaseCO2 emissions because it promotes the bioavailable poolof organic carbon (Zhai et al. 2011). In this study, wefound a trend of higher WSC in the upper layers of FGLsoils compared with FW. In rice paddies, it has beendescribed that nitrogen fertilization increases CH4 emis-sions because it enhances soil carbon inputs decreasesCH4 oxidation due to substrate switch from CH4 to ammo-nia by methanotrophs (Banger et al. 2012). Recently, it hasbeen uncovered that ammonia inhibits the expression ofparticulate CH4 monooxygenase genes in aerobic metha-notrophs (Dam et al. 2014).

Regarding GWP, we found in both types of wetlandsthat CO2 was the main gas contributing to radiative bal-ance during the dry season, while during the rainy andwindy seasons, it was CH4. These results are similar tothose found in restored FW in the southeastern US coastalplain by Morse et al. (2012). In dry wetland areas, theyfound CO2 as the main contributor to the radiative bal-ance, while in flooded wetland areas, the main contributorwas CH4. In this study, during dry season the sum of GWPwas twice higher in FGL than FW, while during rainy andwindy seasons, it was six and five times higher, respec-tively. Hirota et al. (2005) described similar trends indisturbed alpine wetlands in Tibet. The sum of GWP,estimated from CO2 and CH4 fluxes, was 6–11-fold higherunder grazing conditions than under non-grazingconditions.

5. Conclusions

Soil carbon concentration decreased in areas convertedfrom FW to FGL due to decreases in carbon inputs,physical disturbances, and shorter hydroperiods whichenhance higher CO2 and CH4 emissions. However, carbonpools did not decrease in FGL due to an increase in soil

International Journal of Biodiversity Science, Ecosystem Services & Management 9

Dow

nloa

ded

by [

Dic

le U

nive

rsity

] at

08:

03 0

6 N

ovem

ber

2014

BD. Carbon sequestration in wetlands soils is an importantenvironmental service that is negatively affected by chan-ging land use of FW in the flood plains of Veracruz,Mexico. Considering that high CO2 and CH4 emissionsincrease global temperature; if large areas of FW wetlandsare transformed to FGL, then the impacts of these land-usechanges might have repercussions for global warming.Therefore, better policies and law enforcement for fresh-water wetland protection, conservation, and restoration areneeded in Mexico to avoid this positive feedback to globalwarming.

AcknowledgementsFunding for this work was provided by the Mexican NationalCouncil for Science and Technology – CONACYT – throughSector fund CONACYT-SEMARNAT Grant # 107887 and theBasic Science Grant # 081942. The authors thank AlejandroHernández, Monserrat Vidal, J. Alejandro Marín, and CarmeloMaximiliano for their help in the field work. We are also gratefulto the local guides who accompanied us throughout the fieldwork: Tomas León Rodríguez and Eduardo Lauranchet.

ReferencesAltor A, Mitsch WJ. 2006. Methane flux from created riparian

marshes: relationship to intermittent versus continuous inun-dation and emergent macrophytes. Ecol Eng. 28:224–234.doi:10.1016/j.ecoleng.2006.06.006

Badiou P, McDougal R, Pennock D, Clark B. 2011. Greenhousegas emissions and carbon sequestration potential in restoredwetlands of the Canadian prairie pothole region. WetlandsEcol Manage. 19:237–256. doi:10.1007/s11273-011-9214-6

Banger K, Tian H, Lu C. 2012. Do nitrogen fertilizers stimulateor inhibit methane emissions from rice fields? Global ChangeBiol. 18:3259–3267. doi:10.1111/j.1365-2486.2012.02762.x

Bartlett KB, Harriss RC. 1993. Review and assessment of methaneemissions from wetlands. Chemosphere. 26:261–320.doi:10.1016/0045-6535(93)90427-7

Bernal B, Mitsch WJ. 2008. A comparison of soil carbon poolsand profiles in wetlands in Costa Rica and Ohio. Ecol Eng.34:311–323. doi:10.1016/j.ecoleng.2008.09.005

Bernal B, Mitsch WJ. 2012. Comparing carbon sequestration intemperate freshwater wetland communities. Global ChangeBiol. 18:1636–1647. doi:10.1111/j.1365-2486.2011.02619.x

Boeckx P, Cleemput O. 1997. Methane emission from a freshwaterwetland in Belgium. Soil Sci Soc Am J. 61:1250–1256.doi:10.2136/sssaj1997.03615995006100040035x

Bohn HL. 1971. Redox potentials. Soil Sci. 112:39–45.doi:10.1097/00010694-197107000-00007

Brix H, Sorrell BK, Schierup H-H. 1996. Gas fluxes achieved byin situ convective flow in Phragmites Australis. Aquat Bot.54:151–163. doi:10.1016/0304-3770(96)01042-X

Campos A, Hernández ME, Moreno-Casasola P, Cejudo E,Robledo A, Infante D. 2011. Soil water retention and carbonpools in tropical forested wetlands and marshes of the Gulfof Mexico. Hydrolog Sci J. 56:1388–1406. doi:10.1080/02626667.2011.629786

Cerón-Bretón JG, Cerón-Bretón RM, Rangel-Marrón M, Muriel-García M, Cordoba-Quiroz AV, Estrella-Cahuich A. 2011.Determination of carbon sequestration rate in soil of a man-grove forest in Campeche, Mexico. Int J Energ Environ.3:328–336.

Chapelle F, Haack S, Adriens PA, Henry M, Bradley A. 1996.Comparison of Eh and H2 measurements for delineating

redox processes in a contaminated aquifer. Environ SciTechnol. 30:3565–3569. doi:10.1021/es960249+

Cheng L, Leavitt SW, Kimball BA, Pinter Jr PJ, Ottman MJ,Matthias A, Wall GW, Brooks T, Williams DG, Thompson TL.2007. Dynamics of labile and recalcitrant soil carbon pools in asorghum free-air CO2 enrichment agroecosystem. Soil BiolBiochem. 39:2250–2263. doi:10.1016/j.soilbio.2007.03.031

Coles JRP, Yavitt JB. 2004. Linking below ground carbon allo-cation to anaerobic CH4 and CO2 production in a forestedpeatland, New York state. Geomicrobiol J. 21:445–455.doi:10.1080/01490450490505419

Contreras-Espinosa F, Warner BG. 2004. Ecosystem characteris-tics and management considerations for coastal wetlands inMexico. Hydrobiologia. 511:233–245. doi:10.1023/B:HYDR.0000014097.74263.54

Craft CB, Broome SW, Seneca ED. 1988. Nitrogen, phosphorusand organic carbon pools in natural and transplanted marshsoils. Estuaries. 11:272–280. doi:10.2307/1352014

Dam B, Dam S, Kim Y, Liesack W. 2014. Ammonium inducesdifferential expression of methane and nitrogen metabolism-related genes in Methylocystis sp. strain SC2. EnvironMicrobiol. doi:10.1111/1462-2920.12367

De Jong BHJ, Ochoa-Gaona S, Castillo-Santiago MA, Ramirez-Marcial N. 2000. Carbon flux and patterns of land-use/ land-cover change in the Selva Lacandona, Mexico. AMBIO.29(8):504–511.

Dodla SK, Wang JJ, DeLaune R. 2012. Characterization of labileorganic carbon in coastal wetland soils of the MississippiRiver deltaic plain: relationships to carbon functionalities.Sci Total Environ. 435–436:151–158. doi:10.1016/j.scitotenv.2012.06.090

Duan X, Wang X, Ouyang Z. 2009. Influence of common reed(Phragmites australis) on CH4 production and transport in wet-lands: results from single-plant laboratory experiments. WaterAir Soil Poll. 197:185–191. doi:10.1007/s11270-008-9802-0

Elberling B, Askaer L, Jørgensen C, Joensen H, Kühl M, Glud R,Lauritsen F. 2011. Linking soil O2, CO2, and CH4 concen-trations in a wetland soil: implication for CO2 and CH4

fluxes. Environ Sci Technol. 45:3393–3399. doi:10.1021/es103540k

Fearnside PM, Barbosa RI. 1998. Soil carbon changes fromconversion of forest to pasture in Brazilian Amazonia.Forest Ecol Manag. 108:147–166. doi:10.1016/S0378-1127(98)00222-9

Fenchel T, Blackburn TH. 1979. Bacteria and mineral cycling.London: Academic Press.

Frenzel P, Rudolph J. 1998. Methane emission from a wetlandplant: the role of CH4 oxidation in eriophorum. Plant Soil.202:27–32. doi:10.1023/A:1004348929219

Ghani A, Dexter M, Perrott KW. 2003. Hot-water extractablecarbon in soils: a sensitive measurement for determiningimpacts of fertilisation, grazing and cultivation. Soil BiolBiochem. 35:1231–1243. doi:10.1016/S0038-0717(03)00186-X

Girma T, Don P, Asfaw H, Yilma J, Wagnew A. 2007. Effect oflivestock grazing on soil micro-organisms of cracking andself-mulching vertisol. Ethiop Vet J. 11:141–150.

Hansson L, Bronmark C, Anders Nilsson P, Abjornsson K. 2005.Conflicting demands on wetland ecosystem services: nutrientretention, biodiversity or both? Freshwater Biol. 50:705–714.doi:10.1111/j.1365-2427.2005.01352.x

He Y, Xu ZH, Chen CR, Burton J, Ma Q, Ge Y, Xu JM. 2008.Using light fraction and macroaggregate associated organicmatters as early indicators for management-induced changesin soil chemical and biological properties in adjacent nativeand plantation forests of subtropical Australia. Geoderma.147:116–125. doi:10.1016/j.geoderma.2008.08.002

Hernandez ME, Mitsch WJ. 2006. Influence of hydrologic pulses,flooding frequency, and vegetation on nitrous oxide emissions

10 M.E. Hernandez et al.

Dow

nloa

ded

by [

Dic

le U

nive

rsity

] at

08:

03 0

6 N

ovem

ber

2014

from created riparian marshes. Wetlands. 26:862–877.doi:10.1672/0277-5212(2006)26[862:IOHPFF]2.0.CO;2

Hernandez ME, Mitsch WJ. 2007. Denitrification potential andorganic matter as affected by vegetation community, wetlandage, and plant introduction in created wetlands. J EnvironQual. 36:333–342. doi:10.2134/jeq2006.0139

Hirota M, Tang Y, Hu Q, Kato T, Hirata S, Mo W, Cao G, MarikoS. 2005. The potential importance of grazing to the fluxes ofcarbon dioxide and methane in an alpine wetland on theQinghai-Tibetan plateau. Atmos Environ. 39:5255–5259.doi:10.1016/j.atmosenv.2005.05.036

Howe AJ, Rodríguez JF, Saco PM. 2009. Surface evolution andcarbon sequestration in disturbed and undisturbed wetlandsoils of the Hunter estuary, southeast Australia. Estuar CoastShelf S. 84:75–83. doi:10.1016/j.ecss.2009.06.006

Huang GH, Li XZ, Hu YM, Shi Y, Xiao DN. 2005. Methane(CH4) emission from a natural wetland of northern China. JEnviron Sci Heal. 40:1227–1238. doi:10.1081/ESE-200055666

Huo L, Chen Z, Zou Y, Lu X, Guo J, Tang X. 2013. Effect ofZoige alpine wetland degradation on the density and frac-tions of soil organic carbon. Ecol Eng. 51:287–295.doi:10.1016/j.ecoleng.2012.12.020

Infante D, Moreno-Casasola P, Madero-Vega C. 2012. Litterfallof tropical forested wetlands of Veracruz in the coastal flood-plains of the Gulf of Mexico. Aquat Bot. 98:1–11.doi:10.1016/j.aquabot.2011.11.006

IPCC. 2001. Climate change 2001: synthesis report. In: Watson,R.T. and The Core Writing Team editors. A contribution ofworking groups I, II, and III to the third assessment report ofthe intergovernmental panel on climate change. Cambridge(UK) and New York (NY): Cambridge University Press.

Kludze HK, DeLaune RD. 1994. Methane emissions andgrowth of Spartina patens in response to soil redox inten-sity. Soil Sci Soc Am J. 58:1838–1845. doi:10.2136/sssaj1994.03615995005800060037x

Lal R. 2008. Carbon sequestration. Philos Trans R Soc B: BiolSci. 363:815–830. doi:10.1098/rstb.2007.2185

López-Rosas H, Moreno-Casasola P, Mendelssohn I. 2006.Effects of experimental disturbances on a tropical freshwatermarsh invaded by the African grass Echinochloa pyramida-lis. Wetlands. 26:593–604. doi:10.1672/0277-5212(2006)26[593:EOEDOA]2.0.CO;2

Marín-Muñiz JL, Hernández ME, Moreno-Casasola P. 2014.Comparing soil carbon sequestration in coastal freshwaterwetlands with various geomorphic features and plant com-munities in Veracruz, Mexico. Plant Soil. doi:10.1007/s11104-013-2011-7

Mitra SR, Wassmann R, Vlek P. 2005. An appraisal ofglobal wetland area and its organic carbon stock. Curr Sci.88:25–35.

Mitsch WJ, Gosselink JG. 2007. Wetlands. 4th ed. New York(NY): John Wiley and Sons.

Moreno E, Guerrero A, Gutiérrez M, Ortiz C, Palma D. 2002.Los manglares de Tabasco, una reserva natural de carbono.Madera Bosques. 8:115–128.

Moreno-Casasola P. 2008. Los humedales en México, tendenciasy oportunidades. Cuadernos Biodiversidad. 28:10–18.

Moreno-Casasola P, López-Rosas H, Infante D, Peralta LA,Travieso-Bello AC, Warner BG. 2009. Environmental andanthropogenic factors associated with coastal wetland differ-entiation in La Mancha, Veracruz, Mexico. Plant Ecol.200:37–52. doi:10.1007/s11258-008-9400-7

Morse JL, Ardón M, Bernhardt ES. 2012. Greenhouse gas fluxesin southeastern U.S. coastal plain wetlands under contrastingland uses. Ecol Appl. 22:264–280. doi:10.1890/11-0527.1

Murata T, Nguyen ML, Goh KM. 1995. The effects of long-termsuperphosphate application on soil organic matter contentand composition from an intensively managed New

Zealand pasture. Eur J Soil Sci. 46:257–264. doi:10.1111/j.1365-2389.1995.tb01834.x

Nahlik AM, Mitsch WJ. 2010. Methane emissions from createdriverine wetlands. Wetlands. 30:783–793. doi:10.1007/s13157-010-0038-6

Nguyen LM. 2000. Organic matter composition, microbial biomassand microbial activity in gravel-bed constructed wetlands treat-ing farm dairy wastewaters. Ecol Eng. 16:199–221.doi:10.1016/S0925-8574(00)00044-6

Oates L, Jackson R, Allen-Diaz B. 2008. Grazing removaldecreases the magnitude of methane and the variability ofnitrous oxide emissions from spring-fed wetlands of aCalifornia oak savanna. Wetlands Ecol Manage. 16:395–404.doi:10.1007/s11273-007-9076-0

Pennock D, Yates T, Bedard-Haughn A, Phipps K, Farrell R,McDougal R. 2010. Landscape control on N2O and CH4

emissions from freshwater mineral soil wetlands of theCanadian prairie Photole region. Geoderma. 155:308–319.doi:10.1016/j.geoderma.2009.12.015

Peters V, Conrad R. 1995. Methanogenic and other strictly anae-robic bacteria in desert soils and other oxic soils. Appl EnvirMicrobiol. 61:1673–1676.

Roulet NT. 2000. Peatlands, carbon storage, greenhouse gases,and the kyoto protocol: prospects and significance forCanada. Wetlands. 20:605–615. doi:10.1672/0277-5212(2000)020[0605:PCSGGA]2.0.CO;2

Shang ZH, Feng QS, Wu GL, Ren GH, Long RJ. 2013.Grasslandification has significant impacts on soil carbon,nitrogen and phosphorus of alpine wetlands on the Tibetanplateau. Ecol Eng. 58:170–179. doi:10.1016/j.ecoleng.2013.06.035

Sigua G, Coleman S, Albano J. 2009. Beef cattle pasture towetland reconversion: impact on soil organic carbon andphosphorus dynamics. Ecol Eng. 35:1231–1236.doi:10.1016/j.ecoleng.2009.05.004

Silveira ML, Comerford NM, Reddy KR, Cooper WT, El-RifaiH. 2008. Characterization of soil organic carbon pools byacid hydrolysis. Geoderma. 144:405–414. doi:10.1016/j.geoderma.2008.01.002

Sitaula BK, Hansen S, Sitaula J, Bakken LR. 2000. Methaneoxidation potentials and fluxes in agricultural soil: effects offertilization and soil compaction. Biogeochemistry. 48:323–339.doi:10.1023/A:1006262404600

Smith K, Ball T, Conen F, Dobbie K, Massheder J, Rey A. 2003.Exchange of greenhouse gases between soil and atmosphere:interactions of soil physical factors and biological processes. EurJ Soil Sci. 54:779–791. doi:10.1046/j.1351-0754.2003.0567.x

Solomon S, Qin D, Manning M, Alley RB, Berntsen T, BindoffNL, Chen Z, Chidthaisong A, Gregory JM, Hegerl GC, et al.2007. Technical summary. In: Solomon S, Qin D, ManningM, Chen Z, Marquis M, Averyt KB, Tignor M, Miller HL,editors. Climate change 2007: the physical science basis.Contribution of working group I to the fourth assessmentreport of the intergovernmental panel on climate change.Cambridge (UK): Cambridge University Press.

Stevenson FJ. 1982. Humus chemistry. New York (NY):Wiley.

Teuber LM, Hölzel N, Fraser LH. 2013. Livestock grazing inintermountain depressional wetlands – effects on plant stra-tegies, soil characteristics and biomass. Agr Ecosyst Environ.175:21–28. doi:10.1016/j.agee.2013.04.017

Uchida Y, Nishimura S, Akiyama H. 2012. The relationship ofwater-soluble carbon and hot-water-soluble carbon with soilrespiration in agricultural fields. Agr Ecosyst Environ.156:116–122. doi:10.1016/j.agee.2012.05.012

von Lützow M, Leifeld J, Kainz M, Kögel-Knabner I, Munch JC.2002. Indications for soil organic matter quality in soilsunder different management. Geoderma. 105:243–258.doi:10.1016/S0016-7061(01)00106-9

International Journal of Biodiversity Science, Ecosystem Services & Management 11

Dow

nloa

ded

by [

Dic

le U

nive

rsity

] at

08:

03 0

6 N

ovem

ber

2014

Wang K, Jiang H, Zhang X, Zhou G. 2011. Analysis of spatialand temporal variations of carbon dioxide over China usingSCIAMACHY satellite observations during 2003–2005. Int JRem Sens. 32:815–832. doi:10.1080/01431161.2010.517805

Wang S, Tian H, Liu J, Pan S. 2003. Pattern and change of soilorganic carbon storage in China: 1960–1980s. Tellus B.55:416–427. doi:10.1034/j.1600-0889.2003.00039.x

Wang Z, Delaune RD, Patrick Jr WH, Masscheleyn PH. 1993.Soil redox and pH effects on methane production in aflooded rice soil. Soil Sci Soc Am J. 57:382–385.doi:10.2136/sssaj1993.03615995005700020016x

Whalen SC. 2005. Biogeochemistry of methane exchangebetween natural wetlands and the atmosphere. Environ EngSci. 22:73–94. doi:10.1089/ees.2005.22.73

Whiting GJ, Chanton JP. 2001. Greenhouse carbon balance ofwetlands: methane emission versus carbon sequestration.Tellus B. 53:521–528. doi:10.1034/j.1600-0889.2001.530501.x

Wilson JS, Baldwin DS, Rees GN, Wilson BP. 2011. Theeffects of short term inundation on carbon dynamics,microbial community structure and microbial activity infloodplain soil. River Res Appl. 27:213–225. doi:10.1002/rra.1352

Zhai LM, Liu HB, Zhang J, Huang JZ, Wang BR. 2011. Long-term application of organic manure and mineral fertilizer onN2O and CO2 emissions in a red soil from cultivated maize-wheat rotation in China. Agr Sci China. 10:1748–1757.doi:10.1016/S1671-2927(11)60174-0

Zhang JB, Song CC, Yang WY. 2007. Land use effects on thedistribution of labile organic carbon fractions through soilprofiles. Soil Sci Soc Am J. 70:660–667.

Zou XM, Ruan HH, Fu Y, Yang XD, Sha LQ. 2005. Estimatingsoil labile organic carbon and potential turnover rates using asequential fumigation–incubation procedure. Soil BiolBiochem. 37:1923–1928. doi:10.1016/j.soilbio.2005.02.028

12 M.E. Hernandez et al.

Dow

nloa

ded

by [

Dic

le U

nive

rsity

] at

08:

03 0

6 N

ovem

ber

2014