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

Comparing soil carbon pools and carbon gas fluxes in coastal forested wetlands and flooded... International Journal of Biodiversity Science, Ecosystem Services & Management, 2015 Vol. 11, No. 1, 5–16, http://dx.doi.org/10.1080/21513732.2014.925977 Comparing soil carbon pools and carbon gas fluxes in coastal forested wetlands and flooded grasslands in Veracruz, Mexico a, b c a Maria E. Hernandez *, Jose Luis Marín-Muñiz , Patricia Moreno-Casasola and Violeta Vázquez Red de Manejo Biotecnológico de Recursos, Instituto de Ecología, A.C. Carretera Antigua a Coatepec 351, El Haya, Xalapa, Veracruz, Mexico; Centro de Investigaciones Tropicales, Universidad Veracruzana, Casco de la Ex-Hacienda Lucas Martín, Privada de Araucarias S/N. Col. Periodistas, AP. 525, Xalapa, Veracruz, Mexico; Red de Ecología Funcional, Instituto de Ecología, A.C. Carretera Antigua 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 in the soil carbon storage and carbon gaseous emissions. These changes might have important repercussions for global warming. The aim of this study was to investigate whether the conversion of freshwater forested wetlands (FW) to flooded grasslands (FGL) has affected soil carbon cycling. Soil carbon pools and soil organic carbon (SOC) fractions (water-soluble carbon (WSC), hot-water-soluble carbon (HWSC), and HCl/HF soluble carbon (HCl/HF-SC)) were compared between FW and FGL. Additionally, the seasonal dynamic of methane (CH ) and carbon dioxide (CO ) fluxes were monitored in both 4 2 ecosystems located in the coastal plain of Veracruz State Mexico. In FW, soil organic matter (SOM) concentrations were significantly (P ≤ 0.05) higher than FGL. Soil bulk density (BD) was slightly higher in FGL than FW but it was not 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, CO fluxes −2 −1 −2 −1 (26.38 ± 4.45 g m d ) in FGL were significantly higher (P = 0.023) than in FW (14.36 ± 5.77 g m d ). During the rainy and windy seasons, both CH and CO fluxes were significantly higher (P = 0.000 and P = 0.001) in FGL compared 4 2 with 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; ecosystem services 1. Introduction promote the production and release of methane (CH ), a powerful greenhouse gas (GHG). Wetlands are a major Wetlands are the interface between terrestrial and aquatic source of CH in the atmosphere (Whalen 2005)contributing components of the landscape (Mitsch & Gosselink 2007). 23–40% of the annual terrestrial CH emissions and compris- They are widely recognized for providing several ecosystem ing 77–83% of natural sources (IPCC 2001). In addition to services such as flood control, aquifer recharge, and nutrient CH , carbon dioxide (CO ) is produced in wetlands soils 4 2 removal (Hansson et al. 2005). However, their contribution under both aerobic and anaerobic conditions (Smith et al. to the global cycling of atmospheric gases and their impor- 2003; Coles & Yavitt 2004; Elberling et al. 2011). CH tant role as carbon sinks is less recognized. Wetlands, despite emissions to the atmosphere are an environmental concern occupying relatively small areas of the earth’ssurface (2– because global warming potential (GWP) for CH is 25 times 6%), contain a large proportion of the world’s carbon stored than GWP for CO (Solomon et al. 2007). Therefore, small in terrestrial soil reservoirs (Whiting & Chanton 2001;Mitra increase of CH concentration in the atmosphere might have et al. 2005;Lal 2008). Soil organic carbon (SOC) pool is an important impact on global warming. complex; based on resistance to mineralization, it has been Human activities can alter the carbon stocks in wet- divided into labile, intermediate, and recalcitrant organic lands and the exchange of GHG with the atmosphere carbon pools (Cheng et al. 2007). Soil labile and intermediate (Roulet 2000). For example, it has been reported that C pools have a mean residence time of years to several livestock grazing, significantly reduced the above-ground decades while recalcitrant C pools have a mean residence biomass, net primary productivity, and enhanced CH time of hundreds to thousands years (Zou et al. 2005; Cheng emissions in wetlands on the Qinghai-Tibetan plateau in et al. 2007; Silveira et al. 2008). Labile and intermediate China (Hirota et al. 2005). According to Solomon et al. fractions of organic carbon can respond rapidly to environ- (2007), global increases in CO concentration are due mental change; therefore, they are more sensitive indicators primarily to land-use change and fossil-fuel consumption. of the effects of land use than total SOC (von Lützow et al. Based on measured CO fluxes using satellite observations 2002; Zhang et al. 2007;He et al. 2008). and emission inventories in China, Wang et al. (2011) In wetlands ecosystems, flooding conditions not only reported almost five times more CO emissions in slope allow accumulating significant amounts of carbon, but also *Corresponding author. Email: elizabeth.hernadez@inecol.mx Notes: 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. © 2014 Taylor & Francis 6 M.E. Hernandez et al. grasslands compared with swamp soils, considering simi- located from north to south were Estero Dulce lar areas. However, despite the potential importance, few (20º17ʹ53ʺN, 96º52ʹ19ʺW) and Boquilla de Oro studies are available to assess the effects of livestock (19º49ʹ47ʺN, 96º26ʹ59ʺW) (Figure 1). Flooded grasslands grazing on the GHGs emissions in tropical wetland were established in FW 15–20 years ago. Natural wetlands ecosystems. have been transformed to support cattle ranching as eco- In Mexico, wetlands are located mainly on the coast nomic activity. The transformation included cutting native (Contreras-Espinosa & Warner 2004), and these areas are wetland trees to allow the growth of native and introduced among the most transformed ecosystems in the country flood tolerant grasses. Drainage of these areas has not been (Moreno-Casasola 2008). Coastal wetlands in the Mexican performed. However, the introduction of exotic grasses state of Veracruz are threatened mainly by cattle ranching, causes that flooded grasslands experience shorter hydro- petrochemical activities, and urbanization. Because of this, periods than natural wetlands (López-Rosas et al. 2006). most freshwater wetlands show significant changes in their The studied FW are fenced to exclude grazing, while ecology, such as invasion of exotic species, siltation, pol- flooded grasslands experience from moderate (1.8 animals lution, and changes in their hydrology (López-Rosas et al. per ha) to heavy (3 animals per ha) grazing (Girma et al. 2006; Moreno-Casasola et al. 2009). 2007). These areas are grazed from March to early August, In order to assess the impact of human activity on the and afterward animals are moved to uplands because of carbon sequestration service that wetlands provide; the the rainy season. Flooded grasslands are not fertilized with objective of this study was to compare soil organic matter chemicals; and tillage is not performed in these sites. A (SOM), soil carbon pools, and the seasonal carbon gaseous detailed description of the study sites is shown in Table 1. fluxes (CO and CH ) in both coastal freshwater swamps The climate of the coastal plain of the Gulf of Mexico 2 4 and areas that have been converted to flooded grasslands has three seasons: rainy season (July to October), windy (FGL). We hypothesize that carbon pools in flooded grass- season (November to February) that has cold fronts with lands will be lower than in forested wetlands (FW) due to strong winds and rain, and the dry season (March to June). the decrease of carbon inputs and higher mineralization The annual precipitation mean fluctuates between 1200 rates. The opposite will occur with carbon gaseous fluxes, and 1650 mm. The mean annual temperature varies that is, higher fluxes in FGL due to higher carbon between 17°C and 37°C. A detailed description of the mineralization. study sites is shown in Table 1. 2. Materials and methods 2.2. Soil sampling 2.1. Study site In each type of wetland (FW or FGL), three random The study was carried out in two freshwater FW and two sampling plots (1 m ) were established. In these plots, adjacent flooded grasslands, located on the coastal plain of four soil cores (0.48 m deep × 0.05 m diameter) were the Gulf of Mexico in the state of Veracruz. The study sites taken using a Russian peat borer. This borer has thin 97°0'0''W 96°0'0''W 95°0'0''W 0 10204060 80 Km 97°0'0''W 96°0'0''W 95°0'0''W Figure 1. Location of the study sites in the coastal plain of Veracruz, Mexico. 20°0'0''W 20°0'0''W International Journal of Biodiversity Science, Ecosystem Services & Management 7 Table 1. Characteristics of the studied wetlands in the coastal plain of Veracruz, Mexico. No of years of Animals transformation Study site Location Ecosystem type Dominant plant species per ha to FGL Soil type Estero Dulce 20º17ʹ53ʺN 96º52ʹ19ʺW Forested wetland Pachira aquatica Aubl 0 0 Histic (ED) Gleysol Flooded Cynodon plectostachyus 3 More than 15 grassland Cladium jamaicense Boquilla de 19º49ʹ47ʺN, 96º26ʹ59ʺW Forested wetland Ficus insipida and 0 0 Sapric Oro (BO) Pleuranthodendron Histosol lindenii (Turcz.) Sleumer Flooded Cynodon plectostachyus 2 More than 20 grassland Cladium jamaicense Note: Data from previous studies in the area (Campos et al. 2011; Infante et al. 2012). −2 sharp-edged walls, providing a core without compaction, where: soil dry weight (kg m ) = [sampled soil depth] * distortion, or disturbance. Each core was sectioned off [bulk density], and OC = organic carbon content. with a blade at intervals of 4 cm. One of the four cores Total carbon storage to a 44 cm depth was calculated in each sampling plot was used for analysis of bulk density by adding the carbon stored in each one of the soil layers (BD). The soil layers were placed in aluminum pans hav- (Bernal & Mitsch 2012). ing a predetermined dry weight, and transported to the laboratory where they were stored below 4ºC until they were dried in an oven. 2.4. Soluble organic carbon fractions Composite samples were made with three wet soil sam- Extractions of water-soluble carbon (WSC), hot-water- ples taken from the same depth in each sampling plot. Each soluble carbon (HWSC), and HCl/HF soluble carbon of the mixed wet composted samples were packed in con- (HCl/HF-SC) were carried out according to Hernandez tainers and stored below 4ºC until they were dried at room and Mitsch (2007). Soluble organic carbon concentrations temperature and analyzed for carbon content. in each of the extracts were analyzed in a Total Organic Carbon analyzer (Torch, Teledyne Tekmar). 2.3. Soil analysis In the laboratory, wet soil composite samples were mixed 2.5. Gas measurements and flux calculations again to reduce soil heterogeneity and visible residues of Fluxes of CH were measured in situ once every 2 months vegetation were removed. Composite samples were dried starting in August 2010 until February 2012, and the CO at room temperature, pulverized, and sieved (2 mm). For from February 2011 to February 2012, using the closed quantifying the organic matter, approximately 2 g of dried chamber technique (Altor & Mitsch 2006; Hernandez & soil samples were pretreated with 10 M HCl to avoid Mitsch 2006; Nahlik & Mitsch 2010). The closed chamber possible carbonate interferences (Hernandez & Mitsch consisted of two parts: a base and a removable cap, each 2007). After this, SOM was quantified by loss on ignition made of polyvinyl chloride (PVC) pipe (15 cm diameter). at 450ºC for 4 hours (Craft et al. 1988; Bernal & Mitsch The bases were permanently installed in the swamps in 2008). February 2010 and in June 2010 in the flooded grasslands BD was obtained by drying a known volume of sedi- (four chambers in each type of wetland at the two sites, −3 ment at 105ºC (19.64 cm ); then, it was weighed until a n = 8 for each type of wetland). The bases were 30 cm constant weight was reached. The values obtained were high and inserted approximately 5 cm into the wetland −3 used in the formula BD (g cm ) = Mass/volume. soils; the base had an open bottom and a collar, 5 cm from For the purpose of carbon pool calculations, the the top. The removable cap includes a gray butyl sampling organic carbon percentage was calculated as a portion of port and an alcohol-type thermometer in the top. Every organic matter, using Van Bemmelen’s factor (0.58) which time gas fluxes were measured, the cover was put on the has been used for several wetland soils including these base collar, and water was added to ensure a gas-tight seal tropical wetlands (Wang et al. 2003; Hernandez & Mitsch between the base and the cap. Chambers were closed, and 2007; Marín-Muñiz et al. 2014). The carbon pool was every 5 minutes internal gas samples were taken for the −2 calculated in kg C m , according to the following equa- next 45 minutes and the internal temperature registered. tion (Moreno et al. 2002; Cerón-Bretón et al. 2011): Gas samples (25 ml) were taken using 60-ml propylene syringes (TERUMO) having a one-way stopcock (Lieur). KgCm ¼½ soil dry weight ½ OC Gas samples were injected through rubber septa into pre- 8 M.E. Hernandez et al. evacuated 20-ml glass vials. Septa were boiled before use wells, located in each type of wetland at the three study for 30 minutes to eliminate potential gas leaking. All sites. Monitoring wells were made from PVC pipe (13 mm samples were taken between 10:00 h and 16:00 h (local ID), 3 m in length (inserted 1.5 m in the soil), installed in time) and were analyzed within 72 h after collection. each studied wetland (Infante et al. 2012). Gas concentrations were analyzed on a Perkin Elmer Clarus 500 gas chromatograph equipped with a flame ioniza- 2.7.2. Redox potential tion detector (FID) for CH and also equipped with a metha- nizer to detect levels of CO . For sample separation, a Soil redox potential (Eh) was measured within a 30 cm stainless steel column packed with Poropak Q (80/100 diameter around each chamber, at a soil depth between 0 mesh), 6 ft in length, and 2-mm ID was used. The tempera- and 5 cm using a platinum rode and one calomel reference tures for oven, injector, and detector were at 40, 95, and 200° electrode (Corning 476,340), both connected to a digital Cfor CH , and 60, 80, and 350°C for CO ,respectively. 4 2 multimeter. Platinum electrodes were calibrated in situ −1 Nitrogen (7 ml min ) was used as a carrier gas. CH and before every monitoring with quinhydrone (Aldrich) −1 CO were quantified separately. Matheson gas standards 50 mgl in a pH 4.0 buffer solution (Bohn 1971). balanced with N were used to perform calibration curves. All the individual analyzed gas values (ppm CH and CO ) 4 2 were corrected using the ideal gases law (pv = nRT) accord- 2.8. Statistical analysis ing to the formula (Duan et al. 2009; Nahlik & Mitsch 2010): All statistical analyses were performed with SPSS version 18 for Windows. A Kolmogorov–Smirnov test was used m ¼ðÞ c  P  M =ðÞ T  R to check normality. Physical–chemical variables and carbon in soil data fit normal distributions. One-way −3 where: m is the gas concentration (g m ), c the gas analysis of variance (ANOVA) was used to find out 3 −3 concentration by volume (ppm (cm m )), P the atmo- whether the type of ecosystem had an effect on the spheric pressure (assume 1 atm), M the molecular weight average of soil carbon concentration, BD, WSC, HWSC, −1 of gas (g mol ), R the Universal Gas Constant and HCl/HF-SC. These parameters were compared at the (82.0576 (atm.cm )/(mol-K)), and T the air temperature same depth between FW and FGL using a t-test for paired (K) inside the chamber at the time of each sample. samples. To detect differences in carbon pools between The normalized gas concentrations were used to calcu- two types of ecosystems, a t-test was used. Two-way late gas flux rates (Hernandez & Mitsch 2006) according ANOVA with Tukey comparison was used to determine to the following equation: whether climatic season and ecosystem types had an effect in Eh, water levels, and soil temperature. Data for GHGs Fc ¼ðÞ ðÞ d =d ðÞ V=A 1440 c t failed to meet criteria for normal distribution (P < 0.001); therefore, non-parametric statistical tests were used such −2 −1 where: Fc = is the flux rate (mg m d ), (d /d ) = change in c t as Kruskal-Wallis and Mann-Whitney to compare CH gas concentration over the enclosure period, expressed as and CO emissions between the type of ecosystem and −3 −1 3 2 (mg m min ), V the chamber volume (m ), and A the base climatic season. A non-parametric t-test for paired samples chamber soil-surface area (m ), 1440 = minutes by day. was used to compare the emissions of CH or CO 4 2 For each chamber measurement, gas sample concen- between FW and FGL during the same months of sam- tration values were plotted versus sample time. Microsoft pling. A p-value = 0.05 was used to reveal the statistical TM Excel was used to calculate linear regressions on each significance. flux rate. Results were included only if R was greater than 0.85 (Altor & Mitsch 2006). 3. Results 3.1. Soil organic matter concentration, soil bulk 2.6. Conversion to CO -equivalents density, and carbon pools A GWP factor of 25 for CH (Solomon et al. 2007) was The SOM concentration decreased in the studied FW used to convert CH emissions to CO -equivalents for 4 2 −1 soils according to the depth from 382 to 300 g kg comparing their contributions to the global radiative (Figure 2), while in the FGL, SOC decreased from 172 impact. −1 to 97 g kg . When SOM concentration was compared between FW and FGL at the same depth, they were sig- nificantly higher (P ≤ 0.05) in FW, for all depths. Average 2.7. Physical and chemical analysis −1 SOM in FW (284.25 ± 15.2 g kg ) was also significantly 2.7.1. Water level higher (P = 0.001) than average SOM in FGL −1 When surface water was present, the water level was (134.41 ± 6.33 g kg ). −3 recorded using a measuring stick. When no surface water Soil BD in FGL varied from 0.44 to 0.77 g cm , was present, a sensor connected to a multimeter (Steren) increasing with depth, while BD in the FW varied from −3 was used to detect the water level in four monitoring 0.41 to 0.57 g cm ; and BD did not increase with depth. International Journal of Biodiversity Science, Ecosystem Services & Management 9 Organic matter Bulk density (a) (b) –1 –3 (g kg ) (g cm ) 400 0 0.2 0.4 0.6 1 0 100 200 300 0.8 – – X = 134.4 ± 6.3 X = 0.63 ± 0.02 –1 –3 g kg g cm X = 0.49 ± 0.02 X = 284.3 ± 15.2 –3 –1 g cm g kg FGL FW 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 the graph is the mean of six composite cores at 4 cm of depth. Horizontal bars represent standard errors. Although we observed a trend of higher BD in FGL, Organic carbon extracted by HCl/HF decreased sig- when it was compared to FW, no significant differences nificantly with depth in both types of wetlands were found (P ≥ 0.05) neither at any depth, nor in the (P < 0.05). HCl/HF-SC in FW was higher than in −3 average between FGL (0.63 ± 0.02 g cm ) and FW FGL in the entire profile. Average of HCl/HF-SC in −3 −1 (0.49 ± 0.02 g cm ). the entire profile of FGL (0.29 ± 0.05 g C kg ) When the total carbon pools were calculated to 44 cm, was significantly lower (P = 0.006) than in FW −2 −1 the results were 30.63 ± 5.23 kg C m for FGL and (0.41 ± 0.07 g C kg ). −2 28.14 ± 4.87 kg C m for FW, showing no significant difference (P = 0.735). 3.3. Water level, redox potential, and soil temperature dynamics 3.2. Water-soluble C (WSC), hot-water-soluble C Water level in the studied wetlands ranged from −70 to (HWSC), and HCl/HF soluble C 10 cm (Figures 4a–b) in FW and from −45 to 10 cm in In FGL soils, WSC decreased with depth from 0.70 to FGL, without significant differences (Figure 4a; −1 0.27 g C kg , while in FW soils, WSC increased after P= 0.998). When water level values were averaged per −1 17 cm from 0.50 to 0.90 g C kg . In the top layer, WSC each season, significant differences were observed was slightly higher in FGL than FW but not significantly (P= 0.021) with higher values for the rainy (4.86 ± 2.30 different (P = 0.094). Average WSC in the whole soil and 1.18 ± 3.89 cm in FGL and FW, respectively) and the −1 profile of FGL (0.74 ± 0.08 g C kg ) was also not windy seasons (−4.30 ± 3.99 and 3.53 ± 1.79 cm in FGL significantly different (P = 0.094) from FW and FW, respectively), compared with water levels −1 (0.54 ± 0.09 g C kg )(Figure 3). observed during the dry season (−33.75 ± 10.78 cm in HWSC decreased with depth in both ecosystems; FGL FGL and −37.99 ± 24.24 in FW). showed higher HWSC than FW in the whole profile Eh values in the soil oscillated from −37 to 350 mV in except in the deepest layer. The average of HWSC in the FW and from 0 to 337 mV in FGL (Figures 4c–d), and no −1 whole profile of FGL was 1.19 ± 0.19 g C kg , while in significant differences were observed (Figure 4c; −1 FW it was 1.01 ± 0.15 g C kg being not significantly P = 0.613). When Eh values were averaged by season, Eh different (P = 0.309). average in FW decreased from dry (216.26 ± 136.83 mV) to Depth (cm) 10 M.E. Hernandez et al. –1 g C kg 0.0 0.5 1.0 1.5 0.0 1.0 1.5 2.0 2.50.0 0.2 0.4 0.6 0.8 a) b) c) WSC HWSC HCL/HF-SC X = 0.29 ± 0.05 –1 g C kg X = 1.01 ± 0.15 –1 g C kg X = 0.41 ± 0.07 –1 g C kg X = 0.54 ± 0.09 –1 g C kg FGL X = 1.19 ± 0.19 –1 X = 0.74 ± 0.08 g C kg FW –1 g C kg 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) and averaged by climatic season (right); white bars are flooded grasslands, gray bars are FW. Vertical lines on bars and circles values represent standard error, and different letters indicate significant difference. rainy (63.07 ± 42.16 mV) and windy–rainy seasons 209.22 ± 24.45 mV, respectively). In both types of wetlands (170.98 ± 52.13 mV), while in the FGL, Eh values did not Eh values were not significantly different among the sea- decrease (173.03 ± 164.09, 185.12 ± 6.17 and sons (P = 0.695). Depth (cm) International Journal of Biodiversity Science, Ecosystem Services & Management 11 −2 −1 Soil temperatures in the freshwater wetlands (150.14 ± 75.22 and 145.68 ± 30.47 mg m d in FW ranged from 19 to 37°C in both types of wetlands and FGL, respectively; P = 0.224). During the rainy sea- (Figures 4e–f). Soil temperatures in FGL were 1–2°C son, significantly higher CH (P = 0.000) and CO emis- 4 2 higher than in FW during the study period with few excep- sions (P = 0.001) were found in FGL −2 −1 −2 tions. However, no significant differences were found (4349.03 ± 853.46mg m d and 11.82 ± 1.24 g m −1 −2 −1 between the two types of wetlands (P = 0.271). Seasons d , respectively) than in FW (869.01 ± 314.27mg m d −2 −1 had a significant effect (P = 0.020) on soil temperature with and 4.59 ± 1.87 g m d , respectively). Also in the lower temperatures during the windy season (19–21ºC) as windy season, significantly higher CH (P = 0.001) and compared to rainy (27–28ºC) and dry seasons (28–32ºC). CO (P = 0.014) emissions were found in FGL −2 −1 −2 (3912.01 ± 1378.30mg m d and 10.3 ± 8.07 g m −1 −2 −1 d , respectively) than in FW (481.66 ± 324.92mg m d −2 −1 3.4. Methane and carbon dioxide emissions and 2.98 ± 2.77 g m d , respectively). Both CH and CO fluxes were significantly influenced by 4 2 the season (P = 0.0001, P = 0.0001) with higher CH 3.5. Global warming potential emissions during the rainy (August–October) and windy (December–February) seasons compared with the dry sea- We converted seasonal average emission of CH into son (April–June) (Figure 5). For CO , the opposite CO -equivalents to compare its cumulative contributions 2 2 occurred; high fluxes were observed during the months to global radiative balance (Figure 6). During the dry of dry season and low fluxes during rainy and windy season the main component of GHG fluxes was CO seasons. Additionally, both gas emissions were signifi- flux for both types of ecosystems. For rainy and windy cantly affected by the type of ecosystem. During the dry seasons the main component of GHG fluxes was CH for −2 season, average emissions of CO (26.38 ± 4.45 g m d both types of ecosystems. FGL had statistically higher −1 ) were significantly higher (P = 0.023) in FGL compared (P ≤ 0.05) radiative balance in all rainy and windy sea- −2 −1 −2 −1 with FW (14.36 ± 5.77 g m d ), while CH emissions sons (120 and 108 g m d , respectively) than FW (26 −2 −1 were low and similar in both types of ecosystems and15g m d , respectively). On the other hand, during Figure 5. Methane (gray line) and carbon dioxide (black line) emissions in FGL and FW soils. Vertical lines represent standard errors. Figure 6. Total emissions of methane and carbon dioxide expressed as CO -equivalents according to GWP (CH :25 and CO :1; 2 4 2 Solomon et al. 2007). 12 M.E. Hernandez et al. dry season the radiative balance was not significantly despite the latter had higher carbon concentrations but −2 −1) (0.985) different between FGL (30 g m d and FW lower BD (Huo et al. 2013). −2 −1 (18 g m d ). The simplest methods to measure available C sub- strates or labile carbon in both agricultural and wetland soils are WSC and HWSC. Land-use changes cause soil 4. Discussion degradation, and sometimes these carbon fractions are In the areas converted to FGL, SOC was only 47% of the more sensitive than total organic carbon to such degrada- observed in FW. This is probably due to the decrease in tion (Ghani et al. 2003; Dodla et al. 2012; Uchida et al. the carbon inputs to the soils and changes in hydrology. 2012). We measured WSC and HWSC to investigate Litterfall in these tropical FW has been described as high whether the transformation of FW to FGL has affected −1 as 9–15 ton ha per year (Infante et al. 2012) and a large the carbon cycling. However, no clear differences were part of the carbon remains in the flooded soils. However, observed. A trend of higher WSC in the top layer of FGL parts of these FW were transformed to FGL at least was found but in deeper layers FW had higher concentra- 15 years ago (Table 1). The transformation included clear- tions of WSC. The high concentration of WSC in deeper layers in wetland soil is due to leaching of WSC from the ing trees that allowed the growth of native and exotic top layer due to flooding conditions (Dodla et al. 2012). In flood-tolerant grasses to support cattle grazing. Channelization to drain FGL has not been performed in this study, FGL showed shorter hydroperiods which might these sites; however, changes in vegetation cover have have limited the leaching of WSC. Although not statisti- induced shorter hydroperiods in FGL than in FW cally different, HWSC also showed a trend of higher (Figure 4). Even though water levels were not significantly concentrations in the whole profile of FGL than in FW. different, FW were flooded during windy–rainy season HWSC consists of a labile pool of SOM which includes while the FGL experienced some dry periods. Shorter microbial biomass as well as soluble soil carbohydrates hydroperiods might stimulate high C mineralization as it and amines (Ghani et al. 2003). The fact that this type of was confirmed by observed high carbon fluxes in this carbon was higher in FGL and might also due to shorter study. The loss of organic carbon in degraded wetlands hydroperiods that enhance less reduced conditions and due to changing land use has been described in other higher activity of aerobic microorganisms that hydrolyze tropical wetlands. Sigua et al. (2009) found that natural SOM releasing HWSC. We found that HWSC were three −1 wetlands in South Florida had 180 g C kg while altered to four times higher than WSC, and this occurs because pastures (wetlands converted to pastures for 45 years) had hot water dissolves more complex carbon compounds such −1 only 5.4 g C kg and after 6 years of wetland restoration as microbial biomass C, root exudates, amino acids, and C −1 SOC increased to 10.7 g C kg . On the other hand, in bound to soil enzymes. The results of this study are similar temperate wetlands, Shang et al. (2013) described the term to the values found in coastal wetland soils of the ‘grasslandification’ as the process where wetlands are Mississippi River deltaic plain where HWSC was 4–13 drained and converted to pasture with dominant plants of higher than WSC (Dodla et al. 2012). Also, Ghani et al. grasses. They described that grasslandification for 50 years (2003) found in uplands soils that WSC constituted only in Chinese alpine wetlands had reduced vegetation quality approximately 3–6% of HWSC and this type of carbon has and increased the degree of drought and reduced the C, N, been correlated positively with soil respiration (Uchida and P content of soils. Their observations are similar to the et al. 2012). In this study, the results showed a higher content of HWSC in FGL than in FW, and the former findings in transformed FW to FGL for at least 15 years in Veracruz, Mexico. There are few studies of soil carbon in has the higher carbon gaseous fluxes. On the other hand, Mexican wetlands (Campos et al. 2011; Marín-Muñiz the carbon fraction – HCl/HF-SC was significantly higher et al. 2014), and this study is the first report comparing in FW than in FGL. This fraction represents carbon clo- SOC in disturbed wetlands. In other Mexican ecosystems sely associated with soil minerals (Al and Fe) and it is such as the upland tropical forest ‘La selva Lacandona’, considered less available for microorganisms than WSC land-use change to pasture has decreased soil carbon pools and HWSC (Stevenson 1982; Nguyen 2000). Other stu- by approximately 50% (De Jong et al. 2000). In the dies in upland soils have indicated that Al- and Fe-bound Brazilian Amazonia, conversion of forest to pastures over organic matter fractions were subjected to depletion during several decades also caused a decrease of SOC (Fearnside the harvesting and pasturing (Murata et al. 1995). & Barbosa 1998). We found a strong seasonal influence on carbon gas- Despite of low organic carbon content in FGL, carbon eous fluxes in both FW and FGL. When water tables pools were similar to FW. This happens because carbon dropped during the dry season both types of wetlands pools were calculated using SOC content and BD, and the showed low CH emissions and high CO emissions. In 4 2 latter were higher in FGL than the observed in FW. Higher contrast, when soils were flooded (rainy and windy sea- BD in FGL might be caused by compaction due to cattle sons), higher CH emissions and lower CO emissions 4 2 hoof action and shorter hydroperiods (Howe et al. 2009; were observed. This finding agrees with several studies Teuber et al. 2013). Similar results were obtained in Zoige that have described that CH emissions are favored when alpine wetlands in China; degraded wetlands (flooded soils are flooded (Altor & Mitsch 2006; Nahlik & Mitsch 2010; Morse et al. 2012). However, despite both types of meadows) had higher carbon pools than pristine FW International Journal of Biodiversity Science, Ecosystem Services & Management 13 wetlands showed the same seasonal trend in carbon gas- trend was observed by Pennock et al. (2010) in an ephem- eous fluxes, the magnitude of CH and CO fluxes in FGL eral wetland in Saskatchewan, Canada. Badiou et al. 4 2 was higher than in FW. This agrees with results found in (2011) described that the transition period causes the disturbed wetlands. Hirota et al. (2005) found that live- release of a massive pulse of CH due to the fact that the stock grazing stimulated CH emissions from alpine wet- wetland sediment is still saturated and anoxic favoring lands in Tibet, compared with wetlands without grazing. methanogenesis. Additionally, the sediment surface Also, Oates et al. (2008) observed greater CH emissions would warm dramatically as water levels decrease, thereby under grazing conditions in spring-fed wetlands of a increasing rates of methanogenesis which are known to California oak savanna. In this study, the explanation for increase with soil temperature (Bartlett & Harriss 1993). higher CH emissions in FGL compared with FW might Lastly, the decrease in water column depth above the be due to several factors, including the physical soil dis- sediment–water interface would facilitate the transfer of turbances by hoof action of cattle, changes in vegetation CH from the sediments to the atmosphere while reducing cover, different hydroperiods, and changes in the soil’s the potential for CH consumption within the water col- chemistry due to deposition of cattle excreta on soils umn. The more frequent wet and dry cycles in FGL might during grazing. also be responsible for higher CO emissions than FW Cattle disturb soil porosity and break up the stratifi- during the rainy and windy seasons. Wilson et al. (2011) cation of surface and sub-surface water, which contain O found that carbon mineralization and therefore CO emis- 2 2 and methanotrophic bacteria. It has been described that sions increased significantly after flooding occurred in livestock grazing and agricultural practices may have an riparian wetland soils. effect on the soil’s ability to consume CH by altering the Soil fertilization caused by cattle excreta deposition on distribution of pore space, thereby reducing CH diffu- FGL soil is another factor that might explain the higher sion rates through the soil profile, and slowing transport CO and CH emissions in these sites. Studies in uplands 2 4 to sites of CH oxidizing bacteria (Boeckx & Cleemput soils have shown that manure addition to soils increase 1997). Compaction leads to a reduction of aerobic micro- CO emissions because it promotes the bioavailable pool sites and consequently the decrease of CH oxidation by of organic carbon (Zhai et al. 2011). In this study, we oxidizing bacteria (Sitaula et al. 2000). Our study found found a trend of higher WSC in the upper layers of FGL greater compaction in FGL than in FW and is potentially soils compared with FW. In rice paddies, it has been a major influence on reduction of CH consumption. described that nitrogen fertilization increases CH emis- 4 4 Besides these physical factors, the change from native sions because it enhances soil carbon inputs decreases wetland trees to grasses in FGL might also have CH oxidation due to substrate switch from CH to ammo- 4 4 decreased CH oxidation, because wetland plants supply nia by methanotrophs (Banger et al. 2012). Recently, it has oxygen to the rhizosphere, which enhances areas of been uncovered that ammonia inhibits the expression of potential CH oxidation in the soil (Brix et al. 1996; particulate CH monooxygenase genes in aerobic metha- 4 4 Frenzel & Rudolph 1998). notrophs (Dam et al. 2014). Hydrology is one of the factors controlling Eh which Regarding GWP, we found in both types of wetlands influences biogeochemical process in wetlands soils that CO was the main gas contributing to radiative bal- (Mitsch & Gosselink 2007). The sediment Ehs in the ance during the dry season, while during the rainy and studied sites were moderately reducing (−100 to windy seasons, it was CH . These results are similar to 250 mV) (Bohn 1971). Methanogenesis is such an obligate those found in restored FW in the southeastern US coastal anaerobic process that it would not be expected to occur in plain by Morse et al. (2012). In dry wetland areas, they sediments until the Eh is at least −150 mV (Wang et al. found CO as the main contributor to the radiative bal- 1993; Kludze & DeLaune 1994). However, authors such ance, while in flooded wetland areas, the main contributor as Huang et al. (2005) and Wang et al. (1993) also have was CH . In this study, during dry season the sum of GWP found that methanogenic activities are still active at values was twice higher in FGL than FW, while during rainy and close to −100 mV. Chapelle et al. (1996) described that windy seasons, it was six and five times higher, respec- although Eh measurements are easy to do in the field, they tively. Hirota et al. (2005) described similar trends in do not always indicate with accuracy the anoxic biogeo- disturbed alpine wetlands in Tibet. The sum of GWP, chemical process in the soils; and this is one possible estimated from CO and CH fluxes, was 6–11-fold higher 2 4 explanation to the results in this study. under grazing conditions than under non-grazing Paradoxically, higher CH emissions were found in conditions. FGL, which have shorter hydroperiods and although not statistically a distinguishable high Eh compared with FW. 5. Conclusions This might be due to more wet and dry cycles that FGL experience in comparison with FW, especially during rainy Soil carbon concentration decreased in areas converted and windy seasons (Figure 4a). Badiou et al. (2011) from FW to FGL due to decreases in carbon inputs, described that CH emissions in the restored wetlands of physical disturbances, and shorter hydroperiods which the Canadian prairie pothole region increased dramatically enhance higher CO and CH emissions. However, carbon 2 4 just as the wetland basin was becoming dry. The same pools did not decrease in FGL due to an increase in soil 14 M.E. Hernandez et al. redox processes in a contaminated aquifer. Environ Sci BD. Carbon sequestration in wetlands soils is an important Technol. 30:3565–3569. doi:10.1021/es960249+ environmental service that is negatively affected by chan- Cheng L, Leavitt SW, Kimball BA, Pinter Jr PJ, Ottman MJ, ging land use of FW in the flood plains of Veracruz, Matthias A, Wall GW, Brooks T, Williams DG, Thompson TL. Mexico. Considering that high CO and CH emissions 2 4 2007. Dynamics of labile and recalcitrant soil carbon pools in a increase global temperature; if large areas of FW wetlands sorghum free-air CO enrichment agroecosystem. Soil Biol Biochem. 39:2250–2263. doi:10.1016/j.soilbio.2007.03.031 are transformed to FGL, then the impacts of these land-use Coles JRP, Yavitt JB. 2004. Linking below ground carbon allo- changes might have repercussions for global warming. cation to anaerobic CH and CO production in a forested 4 2 Therefore, better policies and law enforcement for fresh- peatland, New York state. Geomicrobiol J. 21:445–455. water wetland protection, conservation, and restoration are doi:10.1080/01490450490505419 needed in Mexico to avoid this positive feedback to global Contreras-Espinosa F, Warner BG. 2004. Ecosystem characteris- tics and management considerations for coastal wetlands in warming. Mexico. Hydrobiologia. 511:233–245. doi:10.1023/B: HYDR.0000014097.74263.54 Craft CB, Broome SW, Seneca ED. 1988. Nitrogen, phosphorus Acknowledgements and organic carbon pools in natural and transplanted marsh Funding for this work was provided by the Mexican National soils. Estuaries. 11:272–280. doi:10.2307/1352014 Council for Science and Technology – CONACYT – through Dam B, Dam S, Kim Y, Liesack W. 2014. Ammonium induces Sector fund CONACYT-SEMARNAT Grant # 107887 and the differential expression of methane and nitrogen metabolism- Basic Science Grant # 081942. The authors thank Alejandro related genes in Methylocystis sp. strain SC2. Environ Hernández, Monserrat Vidal, J. Alejandro Marín, and Carmelo Microbiol. doi:10.1111/1462-2920.12367 Maximiliano for their help in the field work. We are also grateful De Jong BHJ, Ochoa-Gaona S, Castillo-Santiago MA, Ramirez- to the local guides who accompanied us throughout the field Marcial N. 2000. Carbon flux and patterns of land-use/ land- cover change in the Selva Lacandona, Mexico. AMBIO. work: Tomas León Rodríguez and Eduardo Lauranchet. 29(8):504–511. Dodla SK, Wang JJ, DeLaune R. 2012. Characterization of labile organic carbon in coastal wetland soils of the Mississippi References River deltaic plain: relationships to carbon functionalities. Altor A, Mitsch WJ. 2006. Methane flux from created riparian Sci Total Environ. 435–436:151–158. doi:10.1016/j. marshes: relationship to intermittent versus continuous inun- scitotenv.2012.06.090 dation and emergent macrophytes. Ecol Eng. 28:224–234. Duan X, Wang X, Ouyang Z. 2009. Influence of common reed doi:10.1016/j.ecoleng.2006.06.006 (Phragmites australis)on CH production and transport in wet- Badiou P, McDougal R, Pennock D, Clark B. 2011. Greenhouse lands: results from single-plant laboratory experiments. Water gas emissions and carbon sequestration potential in restored Air Soil Poll. 197:185–191. doi:10.1007/s11270-008-9802-0 wetlands of the Canadian prairie pothole region. Wetlands Elberling B, Askaer L, Jørgensen C, Joensen H, Kühl M, Glud R, Ecol Manage. 19:237–256. doi:10.1007/s11273-011-9214-6 Lauritsen F. 2011. Linking soil O ,CO , and CH concen- 2 2 4 Banger K, Tian H, Lu C. 2012. Do nitrogen fertilizers stimulate trations in a wetland soil: implication for CO and CH 2 4 or inhibit methane emissions from rice fields? Global Change fluxes. Environ Sci Technol. 45:3393–3399. doi:10.1021/ Biol. 18:3259–3267. doi:10.1111/j.1365-2486.2012.02762.x es103540k Bartlett KB, Harriss RC. 1993. Review and assessment of methane Fearnside PM, Barbosa RI. 1998. Soil carbon changes from emissions from wetlands. Chemosphere. 26:261–320. conversion of forest to pasture in Brazilian Amazonia. doi:10.1016/0045-6535(93)90427-7 Forest Ecol Manag. 108:147–166. doi:10.1016/S0378- Bernal B, Mitsch WJ. 2008. A comparison of soil carbon pools 1127(98)00222-9 and profiles in wetlands in Costa Rica and Ohio. Ecol Eng. Fenchel T, Blackburn TH. 1979. Bacteria and mineral cycling. 34:311–323. doi:10.1016/j.ecoleng.2008.09.005 London: Academic Press. Bernal B, Mitsch WJ. 2012. Comparing carbon sequestration in Frenzel P, Rudolph J. 1998. Methane emission from a wetland temperate freshwater wetland communities. Global Change plant: the role of CH oxidation in eriophorum. Plant Soil. Biol. 18:1636–1647. doi:10.1111/j.1365-2486.2011.02619.x 202:27–32. doi:10.1023/A:1004348929219 Boeckx P, Cleemput O. 1997. Methane emission from a freshwater Ghani A, Dexter M, Perrott KW. 2003. Hot-water extractable wetland in Belgium. Soil Sci Soc Am J. 61:1250–1256. carbon in soils: a sensitive measurement for determining doi:10.2136/sssaj1997.03615995006100040035x impacts of fertilisation, grazing and cultivation. Soil Biol Bohn HL. 1971. Redox potentials. Soil Sci. 112:39–45. Biochem. 35:1231–1243. doi:10.1016/S0038-0717(03) doi:10.1097/00010694-197107000-00007 00186-X Brix H, Sorrell BK, Schierup H-H. 1996. Gas fluxes achieved by Girma T, Don P, Asfaw H, Yilma J, Wagnew A. 2007. Effect of in situ convective flow in Phragmites Australis. Aquat Bot. livestock grazing on soil micro-organisms of cracking and 54:151–163. doi:10.1016/0304-3770(96)01042-X self-mulching vertisol. Ethiop Vet J. 11:141–150. Campos A, Hernández ME, Moreno-Casasola P, Cejudo E, Hansson L, Bronmark C, Anders Nilsson P, Abjornsson K. 2005. Robledo A, Infante D. 2011. Soil water retention and carbon Conflicting demands on wetland ecosystem services: nutrient pools in tropical forested wetlands and marshes of the Gulf retention, biodiversity or both? Freshwater Biol. 50:705–714. of Mexico. Hydrolog Sci J. 56:1388–1406. doi:10.1080/ doi:10.1111/j.1365-2427.2005.01352.x 02626667.2011.629786 He Y, Xu ZH, Chen CR, Burton J, Ma Q, Ge Y, Xu JM. 2008. Cerón-Bretón JG, Cerón-Bretón RM, Rangel-Marrón M, Muriel- Using light fraction and macroaggregate associated organic García M, Cordoba-Quiroz AV, Estrella-Cahuich A. 2011. matters as early indicators for management-induced changes Determination of carbon sequestration rate in soil of a man- in soil chemical and biological properties in adjacent native grove forest in Campeche, Mexico. Int J Energ Environ. and plantation forests of subtropical Australia. Geoderma. 3:328–336. 147:116–125. doi:10.1016/j.geoderma.2008.08.002 Chapelle F, Haack S, Adriens PA, Henry M, Bradley A. 1996. Hernandez ME, Mitsch WJ. 2006. Influence of hydrologic pulses, Comparison of Eh and H measurements for delineating flooding frequency, and vegetation on nitrous oxide emissions 2 International Journal of Biodiversity Science, Ecosystem Services & Management 15 from created riparian marshes. Wetlands. 26:862–877. Zealand pasture. Eur J Soil Sci. 46:257–264. doi:10.1111/ doi:10.1672/0277-5212(2006)26[862:IOHPFF]2.0.CO;2 j.1365-2389.1995.tb01834.x Hernandez ME, Mitsch WJ. 2007. Denitrification potential and Nahlik AM, Mitsch WJ. 2010. Methane emissions from created organic matter as affected by vegetation community, wetland riverine wetlands. Wetlands. 30:783–793. doi:10.1007/ age, and plant introduction in created wetlands. J Environ s13157-010-0038-6 Qual. 36:333–342. doi:10.2134/jeq2006.0139 Nguyen LM. 2000. Organic matter composition, microbial biomass Hirota M, Tang Y, Hu Q, Kato T, Hirata S, Mo W, Cao G, Mariko and microbial activity in gravel-bed constructed wetlands treat- S. 2005. The potential importance of grazing to the fluxes of ing farm dairy wastewaters. Ecol Eng. 16:199–221. carbon dioxide and methane in an alpine wetland on the doi:10.1016/S0925-8574(00)00044-6 Qinghai-Tibetan plateau. Atmos Environ. 39:5255–5259. Oates L, Jackson R, Allen-Diaz B. 2008. Grazing removal doi:10.1016/j.atmosenv.2005.05.036 decreases the magnitude of methane and the variability of Howe AJ, Rodríguez JF, Saco PM. 2009. Surface evolution and nitrous oxide emissions from spring-fed wetlands of a carbon sequestration in disturbed and undisturbed wetland California oak savanna. Wetlands Ecol Manage. 16:395–404. soils of the Hunter estuary, southeast Australia. Estuar Coast doi:10.1007/s11273-007-9076-0 Shelf S. 84:75–83. doi:10.1016/j.ecss.2009.06.006 Pennock D, Yates T, Bedard-Haughn A, Phipps K, Farrell R, Huang GH, Li XZ, Hu YM, Shi Y, Xiao DN. 2005. Methane McDougal R. 2010. Landscape control on N O and CH 2 4 (CH ) emission from a natural wetland of northern China. J emissions from freshwater mineral soil wetlands of the Environ Sci Heal. 40:1227–1238. doi:10.1081/ESE- Canadian prairie Photole region. Geoderma. 155:308–319. 200055666 doi:10.1016/j.geoderma.2009.12.015 Huo L, Chen Z, Zou Y, Lu X, Guo J, Tang X. 2013. Effect of Peters V, Conrad R. 1995. Methanogenic and other strictly anae- Zoige alpine wetland degradation on the density and frac- robic bacteria in desert soils and other oxic soils. Appl Envir tions of soil organic carbon. Ecol Eng. 51:287–295. Microbiol. 61:1673–1676. doi:10.1016/j.ecoleng.2012.12.020 Roulet NT. 2000. Peatlands, carbon storage, greenhouse gases, Infante D, Moreno-Casasola P, Madero-Vega C. 2012. Litterfall and the kyoto protocol: prospects and significance for of tropical forested wetlands of Veracruz in the coastal flood- Canada. Wetlands. 20:605–615. doi:10.1672/0277-5212 plains of the Gulf of Mexico. Aquat Bot. 98:1–11. (2000)020[0605:PCSGGA]2.0.CO;2 doi:10.1016/j.aquabot.2011.11.006 Shang ZH, Feng QS, Wu GL, Ren GH, Long RJ. 2013. IPCC. 2001. Climate change 2001: synthesis report. In: Watson, Grasslandification has significant impacts on soil carbon, R.T. and The Core Writing Team editors. A contribution of nitrogen and phosphorus of alpine wetlands on the Tibetan working groups I, II, and III to the third assessment report of plateau. Ecol Eng. 58:170–179. doi:10.1016/j. the intergovernmental panel on climate change. Cambridge ecoleng.2013.06.035 (UK) and New York (NY): Cambridge University Press. Sigua G, Coleman S, Albano J. 2009. Beef cattle pasture to Kludze HK, DeLaune RD. 1994. Methane emissions and wetland reconversion: impact on soil organic carbon and growth of Spartina patens in response to soil redox inten- phosphorus dynamics. Ecol Eng. 35:1231–1236. sity. Soil Sci Soc Am J. 58:1838–1845. doi:10.2136/ doi:10.1016/j.ecoleng.2009.05.004 sssaj1994.03615995005800060037x Silveira ML, Comerford NM, Reddy KR, Cooper WT, El-Rifai Lal R. 2008. Carbon sequestration. Philos Trans R Soc B: Biol H. 2008. Characterization of soil organic carbon pools by Sci. 363:815–830. doi:10.1098/rstb.2007.2185 acid hydrolysis. Geoderma. 144:405–414. doi:10.1016/j. López-Rosas H, Moreno-Casasola P, Mendelssohn I. 2006. geoderma.2008.01.002 Effects of experimental disturbances on a tropical freshwater Sitaula BK, Hansen S, Sitaula J, Bakken LR. 2000. Methane marsh invaded by the African grass Echinochloa pyramida- oxidation potentials and fluxes in agricultural soil: effects of lis. Wetlands. 26:593–604. doi:10.1672/0277-5212(2006)26 fertilization and soil compaction. Biogeochemistry. 48:323–339. [593:EOEDOA]2.0.CO;2 doi:10.1023/A:1006262404600 Marín-Muñiz JL, Hernández ME, Moreno-Casasola P. 2014. Smith K, Ball T, Conen F, Dobbie K, Massheder J, Rey A. 2003. Comparing soil carbon sequestration in coastal freshwater Exchange of greenhouse gases between soil and atmosphere: wetlands with various geomorphic features and plant com- interactions of soil physical factors and biological processes. Eur munities in Veracruz, Mexico. Plant Soil. doi:10.1007/ J Soil Sci. 54:779–791. doi:10.1046/j.1351-0754.2003.0567.x s11104-013-2011-7 Solomon S, Qin D, Manning M, Alley RB, Berntsen T, Bindoff Mitra SR, Wassmann R, Vlek P. 2005. An appraisal of NL, Chen Z, Chidthaisong A, Gregory JM, Hegerl GC, et al. global wetland area and its organic carbon stock. Curr Sci. 2007. Technical summary. In: Solomon S, Qin D, Manning 88:25–35. M, Chen Z, Marquis M, Averyt KB, Tignor M, Miller HL, Mitsch WJ, Gosselink JG. 2007. Wetlands. 4th ed. New York editors. Climate change 2007: the physical science basis. (NY): John Wiley and Sons. Contribution of working group I to the fourth assessment Moreno E, Guerrero A, Gutiérrez M, Ortiz C, Palma D. 2002. report of the intergovernmental panel on climate change. Los manglares de Tabasco, una reserva natural de carbono. Cambridge (UK): Cambridge University Press. Madera Bosques. 8:115–128. Stevenson FJ. 1982. Humus chemistry. New York (NY): Moreno-Casasola P. 2008. Los humedales en México, tendencias Wiley. y oportunidades. Cuadernos Biodiversidad. 28:10–18. Teuber LM, Hölzel N, Fraser LH. 2013. Livestock grazing in Moreno-Casasola P, López-Rosas H, Infante D, Peralta LA, intermountain depressional wetlands – effects on plant stra- Travieso-Bello AC, Warner BG. 2009. Environmental and tegies, soil characteristics and biomass. Agr Ecosyst Environ. anthropogenic factors associated with coastal wetland differ- 175:21–28. doi:10.1016/j.agee.2013.04.017 entiation in La Mancha, Veracruz, Mexico. Plant Ecol. Uchida Y, Nishimura S, Akiyama H. 2012. The relationship of 200:37–52. doi:10.1007/s11258-008-9400-7 water-soluble carbon and hot-water-soluble carbon with soil Morse JL, Ardón M, Bernhardt ES. 2012. Greenhouse gas fluxes respiration in agricultural fields. Agr Ecosyst Environ. in southeastern U.S. coastal plain wetlands under contrasting 156:116–122. doi:10.1016/j.agee.2012.05.012 land uses. Ecol Appl. 22:264–280. doi:10.1890/11-0527.1 von Lützow M, Leifeld J, Kainz M, Kögel-Knabner I, Munch JC. Murata T, Nguyen ML, Goh KM. 1995. The effects of long-term 2002. Indications for soil organic matter quality in soils superphosphate application on soil organic matter content under different management. Geoderma. 105:243–258. and composition from an intensively managed New doi:10.1016/S0016-7061(01)00106-9 16 M.E. Hernandez et al. Wilson JS,Baldwin DS,Rees GN,WilsonBP. 2011.The Wang K, Jiang H, Zhang X, Zhou G. 2011. Analysis of spatial effects of short term inundation on carbon dynamics, and temporal variations of carbon dioxide over China using microbial community structure and microbial activity in SCIAMACHY satellite observations during 2003–2005. Int J floodplain soil. River Res Appl. 27:213–225. doi:10.1002/ Rem Sens. 32:815–832. doi:10.1080/01431161.2010.517805 rra.1352 Wang S, Tian H, Liu J, Pan S. 2003. Pattern and change of soil Zhai LM, Liu HB, Zhang J, Huang JZ, Wang BR. 2011. Long- organic carbon storage in China: 1960–1980s. Tellus B. term application of organic manure and mineral fertilizer on 55:416–427. doi:10.1034/j.1600-0889.2003.00039.x N O and CO emissions in a red soil from cultivated maize- Wang Z, Delaune RD, Patrick Jr WH, Masscheleyn PH. 1993. 2 2 wheat rotation in China. Agr Sci China. 10:1748–1757. Soil redox and pH effects on methane production in a doi:10.1016/S1671-2927(11)60174-0 flooded rice soil. Soil Sci Soc Am J. 57:382–385. Zhang JB, Song CC, Yang WY. 2007. Land use effects on the doi:10.2136/sssaj1993.03615995005700020016x distribution of labile organic carbon fractions through soil Whalen SC. 2005. Biogeochemistry of methane exchange profiles. Soil Sci Soc Am J. 70:660–667. between natural wetlands and the atmosphere. Environ Eng Zou XM, Ruan HH, Fu Y, Yang XD, Sha LQ. 2005. Estimating Sci. 22:73–94. doi:10.1089/ees.2005.22.73 Whiting GJ, Chanton JP. 2001. Greenhouse carbon balance of soil labile organic carbon and potential turnover rates using a wetlands: methane emission versus carbon sequestration. sequential fumigation–incubation procedure. Soil Biol Tellus B. 53:521–528. doi:10.1034/j.1600-0889.2001.530501.x Biochem. 37:1923–1928. doi:10.1016/j.soilbio.2005.02.028 http://www.deepdyve.com/assets/images/DeepDyve-Logo-lg.png International Journal of Biodiversity Science, Ecosystem Services & Management Taylor & Francis

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

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2151-3732
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10.1080/21513732.2014.925977
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Abstract

International Journal of Biodiversity Science, Ecosystem Services & Management, 2015 Vol. 11, No. 1, 5–16, http://dx.doi.org/10.1080/21513732.2014.925977 Comparing soil carbon pools and carbon gas fluxes in coastal forested wetlands and flooded grasslands in Veracruz, Mexico a, b c a Maria E. Hernandez *, Jose Luis Marín-Muñiz , Patricia Moreno-Casasola and Violeta Vázquez Red de Manejo Biotecnológico de Recursos, Instituto de Ecología, A.C. Carretera Antigua a Coatepec 351, El Haya, Xalapa, Veracruz, Mexico; Centro de Investigaciones Tropicales, Universidad Veracruzana, Casco de la Ex-Hacienda Lucas Martín, Privada de Araucarias S/N. Col. Periodistas, AP. 525, Xalapa, Veracruz, Mexico; Red de Ecología Funcional, Instituto de Ecología, A.C. Carretera Antigua 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 in the soil carbon storage and carbon gaseous emissions. These changes might have important repercussions for global warming. The aim of this study was to investigate whether the conversion of freshwater forested wetlands (FW) to flooded grasslands (FGL) has affected soil carbon cycling. Soil carbon pools and soil organic carbon (SOC) fractions (water-soluble carbon (WSC), hot-water-soluble carbon (HWSC), and HCl/HF soluble carbon (HCl/HF-SC)) were compared between FW and FGL. Additionally, the seasonal dynamic of methane (CH ) and carbon dioxide (CO ) fluxes were monitored in both 4 2 ecosystems located in the coastal plain of Veracruz State Mexico. In FW, soil organic matter (SOM) concentrations were significantly (P ≤ 0.05) higher than FGL. Soil bulk density (BD) was slightly higher in FGL than FW but it was not 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, CO fluxes −2 −1 −2 −1 (26.38 ± 4.45 g m d ) in FGL were significantly higher (P = 0.023) than in FW (14.36 ± 5.77 g m d ). During the rainy and windy seasons, both CH and CO fluxes were significantly higher (P = 0.000 and P = 0.001) in FGL compared 4 2 with 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; ecosystem services 1. Introduction promote the production and release of methane (CH ), a powerful greenhouse gas (GHG). Wetlands are a major Wetlands are the interface between terrestrial and aquatic source of CH in the atmosphere (Whalen 2005)contributing components of the landscape (Mitsch & Gosselink 2007). 23–40% of the annual terrestrial CH emissions and compris- They are widely recognized for providing several ecosystem ing 77–83% of natural sources (IPCC 2001). In addition to services such as flood control, aquifer recharge, and nutrient CH , carbon dioxide (CO ) is produced in wetlands soils 4 2 removal (Hansson et al. 2005). However, their contribution under both aerobic and anaerobic conditions (Smith et al. to the global cycling of atmospheric gases and their impor- 2003; Coles & Yavitt 2004; Elberling et al. 2011). CH tant role as carbon sinks is less recognized. Wetlands, despite emissions to the atmosphere are an environmental concern occupying relatively small areas of the earth’ssurface (2– because global warming potential (GWP) for CH is 25 times 6%), contain a large proportion of the world’s carbon stored than GWP for CO (Solomon et al. 2007). Therefore, small in terrestrial soil reservoirs (Whiting & Chanton 2001;Mitra increase of CH concentration in the atmosphere might have et al. 2005;Lal 2008). Soil organic carbon (SOC) pool is an important impact on global warming. complex; based on resistance to mineralization, it has been Human activities can alter the carbon stocks in wet- divided into labile, intermediate, and recalcitrant organic lands and the exchange of GHG with the atmosphere carbon pools (Cheng et al. 2007). Soil labile and intermediate (Roulet 2000). For example, it has been reported that C pools have a mean residence time of years to several livestock grazing, significantly reduced the above-ground decades while recalcitrant C pools have a mean residence biomass, net primary productivity, and enhanced CH time of hundreds to thousands years (Zou et al. 2005; Cheng emissions in wetlands on the Qinghai-Tibetan plateau in et al. 2007; Silveira et al. 2008). Labile and intermediate China (Hirota et al. 2005). According to Solomon et al. fractions of organic carbon can respond rapidly to environ- (2007), global increases in CO concentration are due mental change; therefore, they are more sensitive indicators primarily to land-use change and fossil-fuel consumption. of the effects of land use than total SOC (von Lützow et al. Based on measured CO fluxes using satellite observations 2002; Zhang et al. 2007;He et al. 2008). and emission inventories in China, Wang et al. (2011) In wetlands ecosystems, flooding conditions not only reported almost five times more CO emissions in slope allow accumulating significant amounts of carbon, but also *Corresponding author. Email: elizabeth.hernadez@inecol.mx Notes: 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. © 2014 Taylor & Francis 6 M.E. Hernandez et al. grasslands compared with swamp soils, considering simi- located from north to south were Estero Dulce lar areas. However, despite the potential importance, few (20º17ʹ53ʺN, 96º52ʹ19ʺW) and Boquilla de Oro studies are available to assess the effects of livestock (19º49ʹ47ʺN, 96º26ʹ59ʺW) (Figure 1). Flooded grasslands grazing on the GHGs emissions in tropical wetland were established in FW 15–20 years ago. Natural wetlands ecosystems. have been transformed to support cattle ranching as eco- In Mexico, wetlands are located mainly on the coast nomic activity. The transformation included cutting native (Contreras-Espinosa & Warner 2004), and these areas are wetland trees to allow the growth of native and introduced among the most transformed ecosystems in the country flood tolerant grasses. Drainage of these areas has not been (Moreno-Casasola 2008). Coastal wetlands in the Mexican performed. However, the introduction of exotic grasses state of Veracruz are threatened mainly by cattle ranching, causes that flooded grasslands experience shorter hydro- petrochemical activities, and urbanization. Because of this, periods than natural wetlands (López-Rosas et al. 2006). most freshwater wetlands show significant changes in their The studied FW are fenced to exclude grazing, while ecology, such as invasion of exotic species, siltation, pol- flooded grasslands experience from moderate (1.8 animals lution, and changes in their hydrology (López-Rosas et al. per ha) to heavy (3 animals per ha) grazing (Girma et al. 2006; Moreno-Casasola et al. 2009). 2007). These areas are grazed from March to early August, In order to assess the impact of human activity on the and afterward animals are moved to uplands because of carbon sequestration service that wetlands provide; the the rainy season. Flooded grasslands are not fertilized with objective of this study was to compare soil organic matter chemicals; and tillage is not performed in these sites. A (SOM), soil carbon pools, and the seasonal carbon gaseous detailed description of the study sites is shown in Table 1. fluxes (CO and CH ) in both coastal freshwater swamps The climate of the coastal plain of the Gulf of Mexico 2 4 and areas that have been converted to flooded grasslands has three seasons: rainy season (July to October), windy (FGL). We hypothesize that carbon pools in flooded grass- season (November to February) that has cold fronts with lands will be lower than in forested wetlands (FW) due to strong winds and rain, and the dry season (March to June). the decrease of carbon inputs and higher mineralization The annual precipitation mean fluctuates between 1200 rates. The opposite will occur with carbon gaseous fluxes, and 1650 mm. The mean annual temperature varies that is, higher fluxes in FGL due to higher carbon between 17°C and 37°C. A detailed description of the mineralization. study sites is shown in Table 1. 2. Materials and methods 2.2. Soil sampling 2.1. Study site In each type of wetland (FW or FGL), three random The study was carried out in two freshwater FW and two sampling plots (1 m ) were established. In these plots, adjacent flooded grasslands, located on the coastal plain of four soil cores (0.48 m deep × 0.05 m diameter) were the Gulf of Mexico in the state of Veracruz. The study sites taken using a Russian peat borer. This borer has thin 97°0'0''W 96°0'0''W 95°0'0''W 0 10204060 80 Km 97°0'0''W 96°0'0''W 95°0'0''W Figure 1. Location of the study sites in the coastal plain of Veracruz, Mexico. 20°0'0''W 20°0'0''W International Journal of Biodiversity Science, Ecosystem Services & Management 7 Table 1. Characteristics of the studied wetlands in the coastal plain of Veracruz, Mexico. No of years of Animals transformation Study site Location Ecosystem type Dominant plant species per ha to FGL Soil type Estero Dulce 20º17ʹ53ʺN 96º52ʹ19ʺW Forested wetland Pachira aquatica Aubl 0 0 Histic (ED) Gleysol Flooded Cynodon plectostachyus 3 More than 15 grassland Cladium jamaicense Boquilla de 19º49ʹ47ʺN, 96º26ʹ59ʺW Forested wetland Ficus insipida and 0 0 Sapric Oro (BO) Pleuranthodendron Histosol lindenii (Turcz.) Sleumer Flooded Cynodon plectostachyus 2 More than 20 grassland Cladium jamaicense Note: Data from previous studies in the area (Campos et al. 2011; Infante et al. 2012). −2 sharp-edged walls, providing a core without compaction, where: soil dry weight (kg m ) = [sampled soil depth] * distortion, or disturbance. Each core was sectioned off [bulk density], and OC = organic carbon content. with a blade at intervals of 4 cm. One of the four cores Total carbon storage to a 44 cm depth was calculated in each sampling plot was used for analysis of bulk density by adding the carbon stored in each one of the soil layers (BD). The soil layers were placed in aluminum pans hav- (Bernal & Mitsch 2012). ing a predetermined dry weight, and transported to the laboratory where they were stored below 4ºC until they were dried in an oven. 2.4. Soluble organic carbon fractions Composite samples were made with three wet soil sam- Extractions of water-soluble carbon (WSC), hot-water- ples taken from the same depth in each sampling plot. Each soluble carbon (HWSC), and HCl/HF soluble carbon of the mixed wet composted samples were packed in con- (HCl/HF-SC) were carried out according to Hernandez tainers and stored below 4ºC until they were dried at room and Mitsch (2007). Soluble organic carbon concentrations temperature and analyzed for carbon content. in each of the extracts were analyzed in a Total Organic Carbon analyzer (Torch, Teledyne Tekmar). 2.3. Soil analysis In the laboratory, wet soil composite samples were mixed 2.5. Gas measurements and flux calculations again to reduce soil heterogeneity and visible residues of Fluxes of CH were measured in situ once every 2 months vegetation were removed. Composite samples were dried starting in August 2010 until February 2012, and the CO at room temperature, pulverized, and sieved (2 mm). For from February 2011 to February 2012, using the closed quantifying the organic matter, approximately 2 g of dried chamber technique (Altor & Mitsch 2006; Hernandez & soil samples were pretreated with 10 M HCl to avoid Mitsch 2006; Nahlik & Mitsch 2010). The closed chamber possible carbonate interferences (Hernandez & Mitsch consisted of two parts: a base and a removable cap, each 2007). After this, SOM was quantified by loss on ignition made of polyvinyl chloride (PVC) pipe (15 cm diameter). at 450ºC for 4 hours (Craft et al. 1988; Bernal & Mitsch The bases were permanently installed in the swamps in 2008). February 2010 and in June 2010 in the flooded grasslands BD was obtained by drying a known volume of sedi- (four chambers in each type of wetland at the two sites, −3 ment at 105ºC (19.64 cm ); then, it was weighed until a n = 8 for each type of wetland). The bases were 30 cm constant weight was reached. The values obtained were high and inserted approximately 5 cm into the wetland −3 used in the formula BD (g cm ) = Mass/volume. soils; the base had an open bottom and a collar, 5 cm from For the purpose of carbon pool calculations, the the top. The removable cap includes a gray butyl sampling organic carbon percentage was calculated as a portion of port and an alcohol-type thermometer in the top. Every organic matter, using Van Bemmelen’s factor (0.58) which time gas fluxes were measured, the cover was put on the has been used for several wetland soils including these base collar, and water was added to ensure a gas-tight seal tropical wetlands (Wang et al. 2003; Hernandez & Mitsch between the base and the cap. Chambers were closed, and 2007; Marín-Muñiz et al. 2014). The carbon pool was every 5 minutes internal gas samples were taken for the −2 calculated in kg C m , according to the following equa- next 45 minutes and the internal temperature registered. tion (Moreno et al. 2002; Cerón-Bretón et al. 2011): Gas samples (25 ml) were taken using 60-ml propylene syringes (TERUMO) having a one-way stopcock (Lieur). KgCm ¼½ soil dry weight ½ OC Gas samples were injected through rubber septa into pre- 8 M.E. Hernandez et al. evacuated 20-ml glass vials. Septa were boiled before use wells, located in each type of wetland at the three study for 30 minutes to eliminate potential gas leaking. All sites. Monitoring wells were made from PVC pipe (13 mm samples were taken between 10:00 h and 16:00 h (local ID), 3 m in length (inserted 1.5 m in the soil), installed in time) and were analyzed within 72 h after collection. each studied wetland (Infante et al. 2012). Gas concentrations were analyzed on a Perkin Elmer Clarus 500 gas chromatograph equipped with a flame ioniza- 2.7.2. Redox potential tion detector (FID) for CH and also equipped with a metha- nizer to detect levels of CO . For sample separation, a Soil redox potential (Eh) was measured within a 30 cm stainless steel column packed with Poropak Q (80/100 diameter around each chamber, at a soil depth between 0 mesh), 6 ft in length, and 2-mm ID was used. The tempera- and 5 cm using a platinum rode and one calomel reference tures for oven, injector, and detector were at 40, 95, and 200° electrode (Corning 476,340), both connected to a digital Cfor CH , and 60, 80, and 350°C for CO ,respectively. 4 2 multimeter. Platinum electrodes were calibrated in situ −1 Nitrogen (7 ml min ) was used as a carrier gas. CH and before every monitoring with quinhydrone (Aldrich) −1 CO were quantified separately. Matheson gas standards 50 mgl in a pH 4.0 buffer solution (Bohn 1971). balanced with N were used to perform calibration curves. All the individual analyzed gas values (ppm CH and CO ) 4 2 were corrected using the ideal gases law (pv = nRT) accord- 2.8. Statistical analysis ing to the formula (Duan et al. 2009; Nahlik & Mitsch 2010): All statistical analyses were performed with SPSS version 18 for Windows. A Kolmogorov–Smirnov test was used m ¼ðÞ c  P  M =ðÞ T  R to check normality. Physical–chemical variables and carbon in soil data fit normal distributions. One-way −3 where: m is the gas concentration (g m ), c the gas analysis of variance (ANOVA) was used to find out 3 −3 concentration by volume (ppm (cm m )), P the atmo- whether the type of ecosystem had an effect on the spheric pressure (assume 1 atm), M the molecular weight average of soil carbon concentration, BD, WSC, HWSC, −1 of gas (g mol ), R the Universal Gas Constant and HCl/HF-SC. These parameters were compared at the (82.0576 (atm.cm )/(mol-K)), and T the air temperature same depth between FW and FGL using a t-test for paired (K) inside the chamber at the time of each sample. samples. To detect differences in carbon pools between The normalized gas concentrations were used to calcu- two types of ecosystems, a t-test was used. Two-way late gas flux rates (Hernandez & Mitsch 2006) according ANOVA with Tukey comparison was used to determine to the following equation: whether climatic season and ecosystem types had an effect in Eh, water levels, and soil temperature. Data for GHGs Fc ¼ðÞ ðÞ d =d ðÞ V=A 1440 c t failed to meet criteria for normal distribution (P < 0.001); therefore, non-parametric statistical tests were used such −2 −1 where: Fc = is the flux rate (mg m d ), (d /d ) = change in c t as Kruskal-Wallis and Mann-Whitney to compare CH gas concentration over the enclosure period, expressed as and CO emissions between the type of ecosystem and −3 −1 3 2 (mg m min ), V the chamber volume (m ), and A the base climatic season. A non-parametric t-test for paired samples chamber soil-surface area (m ), 1440 = minutes by day. was used to compare the emissions of CH or CO 4 2 For each chamber measurement, gas sample concen- between FW and FGL during the same months of sam- tration values were plotted versus sample time. Microsoft pling. A p-value = 0.05 was used to reveal the statistical TM Excel was used to calculate linear regressions on each significance. flux rate. Results were included only if R was greater than 0.85 (Altor & Mitsch 2006). 3. Results 3.1. Soil organic matter concentration, soil bulk 2.6. Conversion to CO -equivalents density, and carbon pools A GWP factor of 25 for CH (Solomon et al. 2007) was The SOM concentration decreased in the studied FW used to convert CH emissions to CO -equivalents for 4 2 −1 soils according to the depth from 382 to 300 g kg comparing their contributions to the global radiative (Figure 2), while in the FGL, SOC decreased from 172 impact. −1 to 97 g kg . When SOM concentration was compared between FW and FGL at the same depth, they were sig- nificantly higher (P ≤ 0.05) in FW, for all depths. Average 2.7. Physical and chemical analysis −1 SOM in FW (284.25 ± 15.2 g kg ) was also significantly 2.7.1. Water level higher (P = 0.001) than average SOM in FGL −1 When surface water was present, the water level was (134.41 ± 6.33 g kg ). −3 recorded using a measuring stick. When no surface water Soil BD in FGL varied from 0.44 to 0.77 g cm , was present, a sensor connected to a multimeter (Steren) increasing with depth, while BD in the FW varied from −3 was used to detect the water level in four monitoring 0.41 to 0.57 g cm ; and BD did not increase with depth. International Journal of Biodiversity Science, Ecosystem Services & Management 9 Organic matter Bulk density (a) (b) –1 –3 (g kg ) (g cm ) 400 0 0.2 0.4 0.6 1 0 100 200 300 0.8 – – X = 134.4 ± 6.3 X = 0.63 ± 0.02 –1 –3 g kg g cm X = 0.49 ± 0.02 X = 284.3 ± 15.2 –3 –1 g cm g kg FGL FW 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 the graph is the mean of six composite cores at 4 cm of depth. Horizontal bars represent standard errors. Although we observed a trend of higher BD in FGL, Organic carbon extracted by HCl/HF decreased sig- when it was compared to FW, no significant differences nificantly with depth in both types of wetlands were found (P ≥ 0.05) neither at any depth, nor in the (P < 0.05). HCl/HF-SC in FW was higher than in −3 average between FGL (0.63 ± 0.02 g cm ) and FW FGL in the entire profile. Average of HCl/HF-SC in −3 −1 (0.49 ± 0.02 g cm ). the entire profile of FGL (0.29 ± 0.05 g C kg ) When the total carbon pools were calculated to 44 cm, was significantly lower (P = 0.006) than in FW −2 −1 the results were 30.63 ± 5.23 kg C m for FGL and (0.41 ± 0.07 g C kg ). −2 28.14 ± 4.87 kg C m for FW, showing no significant difference (P = 0.735). 3.3. Water level, redox potential, and soil temperature dynamics 3.2. Water-soluble C (WSC), hot-water-soluble C Water level in the studied wetlands ranged from −70 to (HWSC), and HCl/HF soluble C 10 cm (Figures 4a–b) in FW and from −45 to 10 cm in In FGL soils, WSC decreased with depth from 0.70 to FGL, without significant differences (Figure 4a; −1 0.27 g C kg , while in FW soils, WSC increased after P= 0.998). When water level values were averaged per −1 17 cm from 0.50 to 0.90 g C kg . In the top layer, WSC each season, significant differences were observed was slightly higher in FGL than FW but not significantly (P= 0.021) with higher values for the rainy (4.86 ± 2.30 different (P = 0.094). Average WSC in the whole soil and 1.18 ± 3.89 cm in FGL and FW, respectively) and the −1 profile of FGL (0.74 ± 0.08 g C kg ) was also not windy seasons (−4.30 ± 3.99 and 3.53 ± 1.79 cm in FGL significantly different (P = 0.094) from FW and FW, respectively), compared with water levels −1 (0.54 ± 0.09 g C kg )(Figure 3). observed during the dry season (−33.75 ± 10.78 cm in HWSC decreased with depth in both ecosystems; FGL FGL and −37.99 ± 24.24 in FW). showed higher HWSC than FW in the whole profile Eh values in the soil oscillated from −37 to 350 mV in except in the deepest layer. The average of HWSC in the FW and from 0 to 337 mV in FGL (Figures 4c–d), and no −1 whole profile of FGL was 1.19 ± 0.19 g C kg , while in significant differences were observed (Figure 4c; −1 FW it was 1.01 ± 0.15 g C kg being not significantly P = 0.613). When Eh values were averaged by season, Eh different (P = 0.309). average in FW decreased from dry (216.26 ± 136.83 mV) to Depth (cm) 10 M.E. Hernandez et al. –1 g C kg 0.0 0.5 1.0 1.5 0.0 1.0 1.5 2.0 2.50.0 0.2 0.4 0.6 0.8 a) b) c) WSC HWSC HCL/HF-SC X = 0.29 ± 0.05 –1 g C kg X = 1.01 ± 0.15 –1 g C kg X = 0.41 ± 0.07 –1 g C kg X = 0.54 ± 0.09 –1 g C kg FGL X = 1.19 ± 0.19 –1 X = 0.74 ± 0.08 g C kg FW –1 g C kg 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) and averaged by climatic season (right); white bars are flooded grasslands, gray bars are FW. Vertical lines on bars and circles values represent standard error, and different letters indicate significant difference. rainy (63.07 ± 42.16 mV) and windy–rainy seasons 209.22 ± 24.45 mV, respectively). In both types of wetlands (170.98 ± 52.13 mV), while in the FGL, Eh values did not Eh values were not significantly different among the sea- decrease (173.03 ± 164.09, 185.12 ± 6.17 and sons (P = 0.695). Depth (cm) International Journal of Biodiversity Science, Ecosystem Services & Management 11 −2 −1 Soil temperatures in the freshwater wetlands (150.14 ± 75.22 and 145.68 ± 30.47 mg m d in FW ranged from 19 to 37°C in both types of wetlands and FGL, respectively; P = 0.224). During the rainy sea- (Figures 4e–f). Soil temperatures in FGL were 1–2°C son, significantly higher CH (P = 0.000) and CO emis- 4 2 higher than in FW during the study period with few excep- sions (P = 0.001) were found in FGL −2 −1 −2 tions. However, no significant differences were found (4349.03 ± 853.46mg m d and 11.82 ± 1.24 g m −1 −2 −1 between the two types of wetlands (P = 0.271). Seasons d , respectively) than in FW (869.01 ± 314.27mg m d −2 −1 had a significant effect (P = 0.020) on soil temperature with and 4.59 ± 1.87 g m d , respectively). Also in the lower temperatures during the windy season (19–21ºC) as windy season, significantly higher CH (P = 0.001) and compared to rainy (27–28ºC) and dry seasons (28–32ºC). CO (P = 0.014) emissions were found in FGL −2 −1 −2 (3912.01 ± 1378.30mg m d and 10.3 ± 8.07 g m −1 −2 −1 d , respectively) than in FW (481.66 ± 324.92mg m d −2 −1 3.4. Methane and carbon dioxide emissions and 2.98 ± 2.77 g m d , respectively). Both CH and CO fluxes were significantly influenced by 4 2 the season (P = 0.0001, P = 0.0001) with higher CH 3.5. Global warming potential emissions during the rainy (August–October) and windy (December–February) seasons compared with the dry sea- We converted seasonal average emission of CH into son (April–June) (Figure 5). For CO , the opposite CO -equivalents to compare its cumulative contributions 2 2 occurred; high fluxes were observed during the months to global radiative balance (Figure 6). During the dry of dry season and low fluxes during rainy and windy season the main component of GHG fluxes was CO seasons. Additionally, both gas emissions were signifi- flux for both types of ecosystems. For rainy and windy cantly affected by the type of ecosystem. During the dry seasons the main component of GHG fluxes was CH for −2 season, average emissions of CO (26.38 ± 4.45 g m d both types of ecosystems. FGL had statistically higher −1 ) were significantly higher (P = 0.023) in FGL compared (P ≤ 0.05) radiative balance in all rainy and windy sea- −2 −1 −2 −1 with FW (14.36 ± 5.77 g m d ), while CH emissions sons (120 and 108 g m d , respectively) than FW (26 −2 −1 were low and similar in both types of ecosystems and15g m d , respectively). On the other hand, during Figure 5. Methane (gray line) and carbon dioxide (black line) emissions in FGL and FW soils. Vertical lines represent standard errors. Figure 6. Total emissions of methane and carbon dioxide expressed as CO -equivalents according to GWP (CH :25 and CO :1; 2 4 2 Solomon et al. 2007). 12 M.E. Hernandez et al. dry season the radiative balance was not significantly despite the latter had higher carbon concentrations but −2 −1) (0.985) different between FGL (30 g m d and FW lower BD (Huo et al. 2013). −2 −1 (18 g m d ). The simplest methods to measure available C sub- strates or labile carbon in both agricultural and wetland soils are WSC and HWSC. Land-use changes cause soil 4. Discussion degradation, and sometimes these carbon fractions are In the areas converted to FGL, SOC was only 47% of the more sensitive than total organic carbon to such degrada- observed in FW. This is probably due to the decrease in tion (Ghani et al. 2003; Dodla et al. 2012; Uchida et al. the carbon inputs to the soils and changes in hydrology. 2012). We measured WSC and HWSC to investigate Litterfall in these tropical FW has been described as high whether the transformation of FW to FGL has affected −1 as 9–15 ton ha per year (Infante et al. 2012) and a large the carbon cycling. However, no clear differences were part of the carbon remains in the flooded soils. However, observed. A trend of higher WSC in the top layer of FGL parts of these FW were transformed to FGL at least was found but in deeper layers FW had higher concentra- 15 years ago (Table 1). The transformation included clear- tions of WSC. The high concentration of WSC in deeper layers in wetland soil is due to leaching of WSC from the ing trees that allowed the growth of native and exotic top layer due to flooding conditions (Dodla et al. 2012). In flood-tolerant grasses to support cattle grazing. Channelization to drain FGL has not been performed in this study, FGL showed shorter hydroperiods which might these sites; however, changes in vegetation cover have have limited the leaching of WSC. Although not statisti- induced shorter hydroperiods in FGL than in FW cally different, HWSC also showed a trend of higher (Figure 4). Even though water levels were not significantly concentrations in the whole profile of FGL than in FW. different, FW were flooded during windy–rainy season HWSC consists of a labile pool of SOM which includes while the FGL experienced some dry periods. Shorter microbial biomass as well as soluble soil carbohydrates hydroperiods might stimulate high C mineralization as it and amines (Ghani et al. 2003). The fact that this type of was confirmed by observed high carbon fluxes in this carbon was higher in FGL and might also due to shorter study. The loss of organic carbon in degraded wetlands hydroperiods that enhance less reduced conditions and due to changing land use has been described in other higher activity of aerobic microorganisms that hydrolyze tropical wetlands. Sigua et al. (2009) found that natural SOM releasing HWSC. We found that HWSC were three −1 wetlands in South Florida had 180 g C kg while altered to four times higher than WSC, and this occurs because pastures (wetlands converted to pastures for 45 years) had hot water dissolves more complex carbon compounds such −1 only 5.4 g C kg and after 6 years of wetland restoration as microbial biomass C, root exudates, amino acids, and C −1 SOC increased to 10.7 g C kg . On the other hand, in bound to soil enzymes. The results of this study are similar temperate wetlands, Shang et al. (2013) described the term to the values found in coastal wetland soils of the ‘grasslandification’ as the process where wetlands are Mississippi River deltaic plain where HWSC was 4–13 drained and converted to pasture with dominant plants of higher than WSC (Dodla et al. 2012). Also, Ghani et al. grasses. They described that grasslandification for 50 years (2003) found in uplands soils that WSC constituted only in Chinese alpine wetlands had reduced vegetation quality approximately 3–6% of HWSC and this type of carbon has and increased the degree of drought and reduced the C, N, been correlated positively with soil respiration (Uchida and P content of soils. Their observations are similar to the et al. 2012). In this study, the results showed a higher content of HWSC in FGL than in FW, and the former findings in transformed FW to FGL for at least 15 years in Veracruz, Mexico. There are few studies of soil carbon in has the higher carbon gaseous fluxes. On the other hand, Mexican wetlands (Campos et al. 2011; Marín-Muñiz the carbon fraction – HCl/HF-SC was significantly higher et al. 2014), and this study is the first report comparing in FW than in FGL. This fraction represents carbon clo- SOC in disturbed wetlands. In other Mexican ecosystems sely associated with soil minerals (Al and Fe) and it is such as the upland tropical forest ‘La selva Lacandona’, considered less available for microorganisms than WSC land-use change to pasture has decreased soil carbon pools and HWSC (Stevenson 1982; Nguyen 2000). Other stu- by approximately 50% (De Jong et al. 2000). In the dies in upland soils have indicated that Al- and Fe-bound Brazilian Amazonia, conversion of forest to pastures over organic matter fractions were subjected to depletion during several decades also caused a decrease of SOC (Fearnside the harvesting and pasturing (Murata et al. 1995). & Barbosa 1998). We found a strong seasonal influence on carbon gas- Despite of low organic carbon content in FGL, carbon eous fluxes in both FW and FGL. When water tables pools were similar to FW. This happens because carbon dropped during the dry season both types of wetlands pools were calculated using SOC content and BD, and the showed low CH emissions and high CO emissions. In 4 2 latter were higher in FGL than the observed in FW. Higher contrast, when soils were flooded (rainy and windy sea- BD in FGL might be caused by compaction due to cattle sons), higher CH emissions and lower CO emissions 4 2 hoof action and shorter hydroperiods (Howe et al. 2009; were observed. This finding agrees with several studies Teuber et al. 2013). Similar results were obtained in Zoige that have described that CH emissions are favored when alpine wetlands in China; degraded wetlands (flooded soils are flooded (Altor & Mitsch 2006; Nahlik & Mitsch 2010; Morse et al. 2012). However, despite both types of meadows) had higher carbon pools than pristine FW International Journal of Biodiversity Science, Ecosystem Services & Management 13 wetlands showed the same seasonal trend in carbon gas- trend was observed by Pennock et al. (2010) in an ephem- eous fluxes, the magnitude of CH and CO fluxes in FGL eral wetland in Saskatchewan, Canada. Badiou et al. 4 2 was higher than in FW. This agrees with results found in (2011) described that the transition period causes the disturbed wetlands. Hirota et al. (2005) found that live- release of a massive pulse of CH due to the fact that the stock grazing stimulated CH emissions from alpine wet- wetland sediment is still saturated and anoxic favoring lands in Tibet, compared with wetlands without grazing. methanogenesis. Additionally, the sediment surface Also, Oates et al. (2008) observed greater CH emissions would warm dramatically as water levels decrease, thereby under grazing conditions in spring-fed wetlands of a increasing rates of methanogenesis which are known to California oak savanna. In this study, the explanation for increase with soil temperature (Bartlett & Harriss 1993). higher CH emissions in FGL compared with FW might Lastly, the decrease in water column depth above the be due to several factors, including the physical soil dis- sediment–water interface would facilitate the transfer of turbances by hoof action of cattle, changes in vegetation CH from the sediments to the atmosphere while reducing cover, different hydroperiods, and changes in the soil’s the potential for CH consumption within the water col- chemistry due to deposition of cattle excreta on soils umn. The more frequent wet and dry cycles in FGL might during grazing. also be responsible for higher CO emissions than FW Cattle disturb soil porosity and break up the stratifi- during the rainy and windy seasons. Wilson et al. (2011) cation of surface and sub-surface water, which contain O found that carbon mineralization and therefore CO emis- 2 2 and methanotrophic bacteria. It has been described that sions increased significantly after flooding occurred in livestock grazing and agricultural practices may have an riparian wetland soils. effect on the soil’s ability to consume CH by altering the Soil fertilization caused by cattle excreta deposition on distribution of pore space, thereby reducing CH diffu- FGL soil is another factor that might explain the higher sion rates through the soil profile, and slowing transport CO and CH emissions in these sites. Studies in uplands 2 4 to sites of CH oxidizing bacteria (Boeckx & Cleemput soils have shown that manure addition to soils increase 1997). Compaction leads to a reduction of aerobic micro- CO emissions because it promotes the bioavailable pool sites and consequently the decrease of CH oxidation by of organic carbon (Zhai et al. 2011). In this study, we oxidizing bacteria (Sitaula et al. 2000). Our study found found a trend of higher WSC in the upper layers of FGL greater compaction in FGL than in FW and is potentially soils compared with FW. In rice paddies, it has been a major influence on reduction of CH consumption. described that nitrogen fertilization increases CH emis- 4 4 Besides these physical factors, the change from native sions because it enhances soil carbon inputs decreases wetland trees to grasses in FGL might also have CH oxidation due to substrate switch from CH to ammo- 4 4 decreased CH oxidation, because wetland plants supply nia by methanotrophs (Banger et al. 2012). Recently, it has oxygen to the rhizosphere, which enhances areas of been uncovered that ammonia inhibits the expression of potential CH oxidation in the soil (Brix et al. 1996; particulate CH monooxygenase genes in aerobic metha- 4 4 Frenzel & Rudolph 1998). notrophs (Dam et al. 2014). Hydrology is one of the factors controlling Eh which Regarding GWP, we found in both types of wetlands influences biogeochemical process in wetlands soils that CO was the main gas contributing to radiative bal- (Mitsch & Gosselink 2007). The sediment Ehs in the ance during the dry season, while during the rainy and studied sites were moderately reducing (−100 to windy seasons, it was CH . These results are similar to 250 mV) (Bohn 1971). Methanogenesis is such an obligate those found in restored FW in the southeastern US coastal anaerobic process that it would not be expected to occur in plain by Morse et al. (2012). In dry wetland areas, they sediments until the Eh is at least −150 mV (Wang et al. found CO as the main contributor to the radiative bal- 1993; Kludze & DeLaune 1994). However, authors such ance, while in flooded wetland areas, the main contributor as Huang et al. (2005) and Wang et al. (1993) also have was CH . In this study, during dry season the sum of GWP found that methanogenic activities are still active at values was twice higher in FGL than FW, while during rainy and close to −100 mV. Chapelle et al. (1996) described that windy seasons, it was six and five times higher, respec- although Eh measurements are easy to do in the field, they tively. Hirota et al. (2005) described similar trends in do not always indicate with accuracy the anoxic biogeo- disturbed alpine wetlands in Tibet. The sum of GWP, chemical process in the soils; and this is one possible estimated from CO and CH fluxes, was 6–11-fold higher 2 4 explanation to the results in this study. under grazing conditions than under non-grazing Paradoxically, higher CH emissions were found in conditions. FGL, which have shorter hydroperiods and although not statistically a distinguishable high Eh compared with FW. 5. Conclusions This might be due to more wet and dry cycles that FGL experience in comparison with FW, especially during rainy Soil carbon concentration decreased in areas converted and windy seasons (Figure 4a). Badiou et al. (2011) from FW to FGL due to decreases in carbon inputs, described that CH emissions in the restored wetlands of physical disturbances, and shorter hydroperiods which the Canadian prairie pothole region increased dramatically enhance higher CO and CH emissions. However, carbon 2 4 just as the wetland basin was becoming dry. The same pools did not decrease in FGL due to an increase in soil 14 M.E. Hernandez et al. redox processes in a contaminated aquifer. Environ Sci BD. Carbon sequestration in wetlands soils is an important Technol. 30:3565–3569. doi:10.1021/es960249+ environmental service that is negatively affected by chan- Cheng L, Leavitt SW, Kimball BA, Pinter Jr PJ, Ottman MJ, ging land use of FW in the flood plains of Veracruz, Matthias A, Wall GW, Brooks T, Williams DG, Thompson TL. Mexico. Considering that high CO and CH emissions 2 4 2007. Dynamics of labile and recalcitrant soil carbon pools in a increase global temperature; if large areas of FW wetlands sorghum free-air CO enrichment agroecosystem. Soil Biol Biochem. 39:2250–2263. doi:10.1016/j.soilbio.2007.03.031 are transformed to FGL, then the impacts of these land-use Coles JRP, Yavitt JB. 2004. Linking below ground carbon allo- changes might have repercussions for global warming. cation to anaerobic CH and CO production in a forested 4 2 Therefore, better policies and law enforcement for fresh- peatland, New York state. Geomicrobiol J. 21:445–455. water wetland protection, conservation, and restoration are doi:10.1080/01490450490505419 needed in Mexico to avoid this positive feedback to global Contreras-Espinosa F, Warner BG. 2004. Ecosystem characteris- tics and management considerations for coastal wetlands in warming. Mexico. Hydrobiologia. 511:233–245. doi:10.1023/B: HYDR.0000014097.74263.54 Craft CB, Broome SW, Seneca ED. 1988. Nitrogen, phosphorus Acknowledgements and organic carbon pools in natural and transplanted marsh Funding for this work was provided by the Mexican National soils. Estuaries. 11:272–280. doi:10.2307/1352014 Council for Science and Technology – CONACYT – through Dam B, Dam S, Kim Y, Liesack W. 2014. Ammonium induces Sector fund CONACYT-SEMARNAT Grant # 107887 and the differential expression of methane and nitrogen metabolism- Basic Science Grant # 081942. The authors thank Alejandro related genes in Methylocystis sp. strain SC2. Environ Hernández, Monserrat Vidal, J. Alejandro Marín, and Carmelo Microbiol. doi:10.1111/1462-2920.12367 Maximiliano for their help in the field work. We are also grateful De Jong BHJ, Ochoa-Gaona S, Castillo-Santiago MA, Ramirez- to the local guides who accompanied us throughout the field Marcial N. 2000. Carbon flux and patterns of land-use/ land- cover change in the Selva Lacandona, Mexico. AMBIO. work: Tomas León Rodríguez and Eduardo Lauranchet. 29(8):504–511. Dodla SK, Wang JJ, DeLaune R. 2012. Characterization of labile organic carbon in coastal wetland soils of the Mississippi References River deltaic plain: relationships to carbon functionalities. Altor A, Mitsch WJ. 2006. Methane flux from created riparian Sci Total Environ. 435–436:151–158. doi:10.1016/j. marshes: relationship to intermittent versus continuous inun- scitotenv.2012.06.090 dation and emergent macrophytes. Ecol Eng. 28:224–234. Duan X, Wang X, Ouyang Z. 2009. Influence of common reed doi:10.1016/j.ecoleng.2006.06.006 (Phragmites australis)on CH production and transport in wet- Badiou P, McDougal R, Pennock D, Clark B. 2011. Greenhouse lands: results from single-plant laboratory experiments. Water gas emissions and carbon sequestration potential in restored Air Soil Poll. 197:185–191. doi:10.1007/s11270-008-9802-0 wetlands of the Canadian prairie pothole region. Wetlands Elberling B, Askaer L, Jørgensen C, Joensen H, Kühl M, Glud R, Ecol Manage. 19:237–256. doi:10.1007/s11273-011-9214-6 Lauritsen F. 2011. Linking soil O ,CO , and CH concen- 2 2 4 Banger K, Tian H, Lu C. 2012. Do nitrogen fertilizers stimulate trations in a wetland soil: implication for CO and CH 2 4 or inhibit methane emissions from rice fields? Global Change fluxes. Environ Sci Technol. 45:3393–3399. doi:10.1021/ Biol. 18:3259–3267. doi:10.1111/j.1365-2486.2012.02762.x es103540k Bartlett KB, Harriss RC. 1993. Review and assessment of methane Fearnside PM, Barbosa RI. 1998. Soil carbon changes from emissions from wetlands. Chemosphere. 26:261–320. conversion of forest to pasture in Brazilian Amazonia. doi:10.1016/0045-6535(93)90427-7 Forest Ecol Manag. 108:147–166. doi:10.1016/S0378- Bernal B, Mitsch WJ. 2008. A comparison of soil carbon pools 1127(98)00222-9 and profiles in wetlands in Costa Rica and Ohio. Ecol Eng. Fenchel T, Blackburn TH. 1979. Bacteria and mineral cycling. 34:311–323. doi:10.1016/j.ecoleng.2008.09.005 London: Academic Press. Bernal B, Mitsch WJ. 2012. Comparing carbon sequestration in Frenzel P, Rudolph J. 1998. Methane emission from a wetland temperate freshwater wetland communities. Global Change plant: the role of CH oxidation in eriophorum. Plant Soil. Biol. 18:1636–1647. doi:10.1111/j.1365-2486.2011.02619.x 202:27–32. doi:10.1023/A:1004348929219 Boeckx P, Cleemput O. 1997. Methane emission from a freshwater Ghani A, Dexter M, Perrott KW. 2003. Hot-water extractable wetland in Belgium. Soil Sci Soc Am J. 61:1250–1256. carbon in soils: a sensitive measurement for determining doi:10.2136/sssaj1997.03615995006100040035x impacts of fertilisation, grazing and cultivation. Soil Biol Bohn HL. 1971. Redox potentials. Soil Sci. 112:39–45. Biochem. 35:1231–1243. doi:10.1016/S0038-0717(03) doi:10.1097/00010694-197107000-00007 00186-X Brix H, Sorrell BK, Schierup H-H. 1996. Gas fluxes achieved by Girma T, Don P, Asfaw H, Yilma J, Wagnew A. 2007. Effect of in situ convective flow in Phragmites Australis. Aquat Bot. livestock grazing on soil micro-organisms of cracking and 54:151–163. doi:10.1016/0304-3770(96)01042-X self-mulching vertisol. Ethiop Vet J. 11:141–150. Campos A, Hernández ME, Moreno-Casasola P, Cejudo E, Hansson L, Bronmark C, Anders Nilsson P, Abjornsson K. 2005. Robledo A, Infante D. 2011. Soil water retention and carbon Conflicting demands on wetland ecosystem services: nutrient pools in tropical forested wetlands and marshes of the Gulf retention, biodiversity or both? Freshwater Biol. 50:705–714. of Mexico. Hydrolog Sci J. 56:1388–1406. doi:10.1080/ doi:10.1111/j.1365-2427.2005.01352.x 02626667.2011.629786 He Y, Xu ZH, Chen CR, Burton J, Ma Q, Ge Y, Xu JM. 2008. Cerón-Bretón JG, Cerón-Bretón RM, Rangel-Marrón M, Muriel- Using light fraction and macroaggregate associated organic García M, Cordoba-Quiroz AV, Estrella-Cahuich A. 2011. matters as early indicators for management-induced changes Determination of carbon sequestration rate in soil of a man- in soil chemical and biological properties in adjacent native grove forest in Campeche, Mexico. Int J Energ Environ. and plantation forests of subtropical Australia. Geoderma. 3:328–336. 147:116–125. doi:10.1016/j.geoderma.2008.08.002 Chapelle F, Haack S, Adriens PA, Henry M, Bradley A. 1996. Hernandez ME, Mitsch WJ. 2006. Influence of hydrologic pulses, Comparison of Eh and H measurements for delineating flooding frequency, and vegetation on nitrous oxide emissions 2 International Journal of Biodiversity Science, Ecosystem Services & Management 15 from created riparian marshes. Wetlands. 26:862–877. Zealand pasture. Eur J Soil Sci. 46:257–264. doi:10.1111/ doi:10.1672/0277-5212(2006)26[862:IOHPFF]2.0.CO;2 j.1365-2389.1995.tb01834.x Hernandez ME, Mitsch WJ. 2007. Denitrification potential and Nahlik AM, Mitsch WJ. 2010. Methane emissions from created organic matter as affected by vegetation community, wetland riverine wetlands. Wetlands. 30:783–793. doi:10.1007/ age, and plant introduction in created wetlands. J Environ s13157-010-0038-6 Qual. 36:333–342. doi:10.2134/jeq2006.0139 Nguyen LM. 2000. Organic matter composition, microbial biomass Hirota M, Tang Y, Hu Q, Kato T, Hirata S, Mo W, Cao G, Mariko and microbial activity in gravel-bed constructed wetlands treat- S. 2005. The potential importance of grazing to the fluxes of ing farm dairy wastewaters. Ecol Eng. 16:199–221. carbon dioxide and methane in an alpine wetland on the doi:10.1016/S0925-8574(00)00044-6 Qinghai-Tibetan plateau. Atmos Environ. 39:5255–5259. Oates L, Jackson R, Allen-Diaz B. 2008. Grazing removal doi:10.1016/j.atmosenv.2005.05.036 decreases the magnitude of methane and the variability of Howe AJ, Rodríguez JF, Saco PM. 2009. Surface evolution and nitrous oxide emissions from spring-fed wetlands of a carbon sequestration in disturbed and undisturbed wetland California oak savanna. Wetlands Ecol Manage. 16:395–404. soils of the Hunter estuary, southeast Australia. Estuar Coast doi:10.1007/s11273-007-9076-0 Shelf S. 84:75–83. doi:10.1016/j.ecss.2009.06.006 Pennock D, Yates T, Bedard-Haughn A, Phipps K, Farrell R, Huang GH, Li XZ, Hu YM, Shi Y, Xiao DN. 2005. Methane McDougal R. 2010. Landscape control on N O and CH 2 4 (CH ) emission from a natural wetland of northern China. J emissions from freshwater mineral soil wetlands of the Environ Sci Heal. 40:1227–1238. doi:10.1081/ESE- Canadian prairie Photole region. Geoderma. 155:308–319. 200055666 doi:10.1016/j.geoderma.2009.12.015 Huo L, Chen Z, Zou Y, Lu X, Guo J, Tang X. 2013. Effect of Peters V, Conrad R. 1995. Methanogenic and other strictly anae- Zoige alpine wetland degradation on the density and frac- robic bacteria in desert soils and other oxic soils. Appl Envir tions of soil organic carbon. Ecol Eng. 51:287–295. Microbiol. 61:1673–1676. doi:10.1016/j.ecoleng.2012.12.020 Roulet NT. 2000. Peatlands, carbon storage, greenhouse gases, Infante D, Moreno-Casasola P, Madero-Vega C. 2012. Litterfall and the kyoto protocol: prospects and significance for of tropical forested wetlands of Veracruz in the coastal flood- Canada. Wetlands. 20:605–615. doi:10.1672/0277-5212 plains of the Gulf of Mexico. Aquat Bot. 98:1–11. (2000)020[0605:PCSGGA]2.0.CO;2 doi:10.1016/j.aquabot.2011.11.006 Shang ZH, Feng QS, Wu GL, Ren GH, Long RJ. 2013. IPCC. 2001. Climate change 2001: synthesis report. In: Watson, Grasslandification has significant impacts on soil carbon, R.T. and The Core Writing Team editors. A contribution of nitrogen and phosphorus of alpine wetlands on the Tibetan working groups I, II, and III to the third assessment report of plateau. Ecol Eng. 58:170–179. doi:10.1016/j. the intergovernmental panel on climate change. Cambridge ecoleng.2013.06.035 (UK) and New York (NY): Cambridge University Press. Sigua G, Coleman S, Albano J. 2009. Beef cattle pasture to Kludze HK, DeLaune RD. 1994. Methane emissions and wetland reconversion: impact on soil organic carbon and growth of Spartina patens in response to soil redox inten- phosphorus dynamics. Ecol Eng. 35:1231–1236. sity. Soil Sci Soc Am J. 58:1838–1845. doi:10.2136/ doi:10.1016/j.ecoleng.2009.05.004 sssaj1994.03615995005800060037x Silveira ML, Comerford NM, Reddy KR, Cooper WT, El-Rifai Lal R. 2008. Carbon sequestration. Philos Trans R Soc B: Biol H. 2008. Characterization of soil organic carbon pools by Sci. 363:815–830. doi:10.1098/rstb.2007.2185 acid hydrolysis. Geoderma. 144:405–414. doi:10.1016/j. López-Rosas H, Moreno-Casasola P, Mendelssohn I. 2006. geoderma.2008.01.002 Effects of experimental disturbances on a tropical freshwater Sitaula BK, Hansen S, Sitaula J, Bakken LR. 2000. Methane marsh invaded by the African grass Echinochloa pyramida- oxidation potentials and fluxes in agricultural soil: effects of lis. Wetlands. 26:593–604. doi:10.1672/0277-5212(2006)26 fertilization and soil compaction. Biogeochemistry. 48:323–339. [593:EOEDOA]2.0.CO;2 doi:10.1023/A:1006262404600 Marín-Muñiz JL, Hernández ME, Moreno-Casasola P. 2014. Smith K, Ball T, Conen F, Dobbie K, Massheder J, Rey A. 2003. Comparing soil carbon sequestration in coastal freshwater Exchange of greenhouse gases between soil and atmosphere: wetlands with various geomorphic features and plant com- interactions of soil physical factors and biological processes. Eur munities in Veracruz, Mexico. Plant Soil. doi:10.1007/ J Soil Sci. 54:779–791. doi:10.1046/j.1351-0754.2003.0567.x s11104-013-2011-7 Solomon S, Qin D, Manning M, Alley RB, Berntsen T, Bindoff Mitra SR, Wassmann R, Vlek P. 2005. An appraisal of NL, Chen Z, Chidthaisong A, Gregory JM, Hegerl GC, et al. global wetland area and its organic carbon stock. Curr Sci. 2007. Technical summary. In: Solomon S, Qin D, Manning 88:25–35. M, Chen Z, Marquis M, Averyt KB, Tignor M, Miller HL, Mitsch WJ, Gosselink JG. 2007. Wetlands. 4th ed. New York editors. Climate change 2007: the physical science basis. (NY): John Wiley and Sons. Contribution of working group I to the fourth assessment Moreno E, Guerrero A, Gutiérrez M, Ortiz C, Palma D. 2002. report of the intergovernmental panel on climate change. Los manglares de Tabasco, una reserva natural de carbono. Cambridge (UK): Cambridge University Press. Madera Bosques. 8:115–128. Stevenson FJ. 1982. Humus chemistry. New York (NY): Moreno-Casasola P. 2008. Los humedales en México, tendencias Wiley. y oportunidades. Cuadernos Biodiversidad. 28:10–18. Teuber LM, Hölzel N, Fraser LH. 2013. Livestock grazing in Moreno-Casasola P, López-Rosas H, Infante D, Peralta LA, intermountain depressional wetlands – effects on plant stra- Travieso-Bello AC, Warner BG. 2009. Environmental and tegies, soil characteristics and biomass. Agr Ecosyst Environ. anthropogenic factors associated with coastal wetland differ- 175:21–28. doi:10.1016/j.agee.2013.04.017 entiation in La Mancha, Veracruz, Mexico. Plant Ecol. Uchida Y, Nishimura S, Akiyama H. 2012. The relationship of 200:37–52. doi:10.1007/s11258-008-9400-7 water-soluble carbon and hot-water-soluble carbon with soil Morse JL, Ardón M, Bernhardt ES. 2012. Greenhouse gas fluxes respiration in agricultural fields. Agr Ecosyst Environ. in southeastern U.S. coastal plain wetlands under contrasting 156:116–122. doi:10.1016/j.agee.2012.05.012 land uses. Ecol Appl. 22:264–280. doi:10.1890/11-0527.1 von Lützow M, Leifeld J, Kainz M, Kögel-Knabner I, Munch JC. Murata T, Nguyen ML, Goh KM. 1995. The effects of long-term 2002. Indications for soil organic matter quality in soils superphosphate application on soil organic matter content under different management. Geoderma. 105:243–258. and composition from an intensively managed New doi:10.1016/S0016-7061(01)00106-9 16 M.E. Hernandez et al. Wilson JS,Baldwin DS,Rees GN,WilsonBP. 2011.The Wang K, Jiang H, Zhang X, Zhou G. 2011. Analysis of spatial effects of short term inundation on carbon dynamics, and temporal variations of carbon dioxide over China using microbial community structure and microbial activity in SCIAMACHY satellite observations during 2003–2005. Int J floodplain soil. River Res Appl. 27:213–225. doi:10.1002/ Rem Sens. 32:815–832. doi:10.1080/01431161.2010.517805 rra.1352 Wang S, Tian H, Liu J, Pan S. 2003. Pattern and change of soil Zhai LM, Liu HB, Zhang J, Huang JZ, Wang BR. 2011. Long- organic carbon storage in China: 1960–1980s. Tellus B. term application of organic manure and mineral fertilizer on 55:416–427. doi:10.1034/j.1600-0889.2003.00039.x N O and CO emissions in a red soil from cultivated maize- Wang Z, Delaune RD, Patrick Jr WH, Masscheleyn PH. 1993. 2 2 wheat rotation in China. Agr Sci China. 10:1748–1757. Soil redox and pH effects on methane production in a doi:10.1016/S1671-2927(11)60174-0 flooded rice soil. Soil Sci Soc Am J. 57:382–385. Zhang JB, Song CC, Yang WY. 2007. Land use effects on the doi:10.2136/sssaj1993.03615995005700020016x distribution of labile organic carbon fractions through soil Whalen SC. 2005. Biogeochemistry of methane exchange profiles. Soil Sci Soc Am J. 70:660–667. between natural wetlands and the atmosphere. Environ Eng Zou XM, Ruan HH, Fu Y, Yang XD, Sha LQ. 2005. Estimating Sci. 22:73–94. doi:10.1089/ees.2005.22.73 Whiting GJ, Chanton JP. 2001. Greenhouse carbon balance of soil labile organic carbon and potential turnover rates using a wetlands: methane emission versus carbon sequestration. sequential fumigation–incubation procedure. Soil Biol Tellus B. 53:521–528. doi:10.1034/j.1600-0889.2001.530501.x Biochem. 37:1923–1928. doi:10.1016/j.soilbio.2005.02.028

Journal

International Journal of Biodiversity Science, Ecosystem Services & ManagementTaylor & Francis

Published: Jan 2, 2015

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

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