Landscape controls on N2O and CH4 emissions from freshwater mineral soil wetlands of the Canadian Prairie Pothole region

Geoderma 155 (2010) 308–319

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Geoderma j o u r n a l h o m e p a g e : w w w. e l s ev i e r. c o m / l o c a t e / g e o d e r m a

Landscape controls on N2O and CH4 emissions from freshwater mineral soil wetlands of the Canadian Prairie Pothole region Dan Pennock a,⁎, Thomas Yates a, Angela Bedard-Haughn a, Kim Phipps a, Richard Farrell a, Rhonda McDougal b a b

Department of Soil Science, University of Saskatchewan, 51 Campus Drive, Saskatoon, Saskatchewan, Canada S7N 5A8 Manitoba Water Stewardship, 200 Salteaux Crescent, Winnipeg, MB R3J 3W3; formerly Ducks Unlimited Canada, P. O. Box 1150, Stonewall, MB, Canada R0C 2Z0

a r t i c l e

i n f o

Article history: Received 11 June 2009 Received in revised form 30 November 2009 Accepted 14 December 2009 Available online 27 January 2010 Keywords: Nitrous oxide Methane Greenhouse gas Grassland Topography Land use Denitrification Nitrification

a b s t r a c t The characteristic feature of the Prairie Pothole Region is a complex assemblage of mineral soil wetlands embedded in the dominantly agricultural landscape. Soils in these wetlands are loci of high potential greenhouse gas (GHG) emissions, and our objective was to provide estimates of greenhouse gas emissions and the controls on these emissions for typical wetlands of this region. Three years (2004–06) of N2O and CH4 emissions were taken from a large semi-permanent pond and five ephemeral freshwater mineral soil wetlands at the St. Denis National Wildlife Area (SDNWA) near Saskatoon, Saskatchewan, Canada. Methane emissions from the semi-permanent pond were low (ranging from 0.04 to 3.33 g CH4 m− 2 yr-1) but emissions from landscape elements of the ephemeral ponds were substantially higher, with a maximum of 138.6 g CH4 m− 2 yr− 1 (or approximately 110 g CH4 m− 2 yr− 1 when corrected for mid-day sampling bias) from basin centers of these ponds in 2005. The average annual CH4 emissions averaged across the three elements of the ephemeral ponds at SDNWA were 54.8 g CH4 m− 2 yr− 1 in 2005 and 32.4 g CH4 m− 2 yr− 1 in 2006. Methane emissions were significantly inversely correlated to SO24 concentrations of the pond water, which are in turn related to the balance between surface runoff and groundwater inputs into the ponds. The semi-permanent pond consistently had low annual N2O emissions (b 0.4 kg N2O–N ha− 1 yr− 1). N2O emissions from landscape elements within the ephemeral ponds showed considerable inter-annual variation, ranging from 0.09 to 1.0 kg N2O–N ha− 1 yr− 1 for riparian grass elements, 0.3 to 0.6 kg N2O–N ha− 1 yr− 1 for riparian tree, and 1.0 to 2.1 kg N2O–N ha− 1 yr− 1 for basin centers. Major N2O emission events in the wetland elements were associated with periods of rapid drainage (i.e., from greater than 80% to less than 60% water-filled pore space) in the upper 15 cm of the soil. Within-year patterns of N2O and CH4 emissions from soils of the ephemeral ponds were closely related to a second hydrological control, the area and duration of inundation in the ponds but negligible differences were observed between riparian grass and tree elements. The strong interactions between hydrology, water chemistry, and emissions of N2O and CH4 demonstrate the need for a landscape-scale assessment of GHG processes in these landscapes. © 2010 Elsevier B.V. All rights reserved.

1. Introduction Agricultural landscapes of the Prairie Pothole region feature a complex mosaic of land uses and landforms. The dominant geomorphic surface of the Prairie Pothole region is hummocky, composed of chaotically arranged knolls and depressions that formed during melt of stagnant ice masses during the retreat of the Wisconsinan ice sheet that began approximately 17,000 yrs B.P. in southern Saskatchewan. When uncultivated, the depressions are commonly occupied by wetlands and the density of wetlands can be very high; for example, the

⁎ Corresponding author. E-mail address: [email protected] (D. Pennock). 0016-7061/$ – see front matter © 2010 Elsevier B.V. All rights reserved. doi:10.1016/j.geoderma.2009.12.015

362 ha of our research site at St. Denis National Wildlife Area (SDNWA) in Saskatchewan, Canada feature over 200 wetlands (Waiser, 2006). Wetlands of the Prairie Pothole region are classified as freshwater mineral soil wetlands (Bridgham et al., 2006) rather than organic as in peat-dominated wetlands. The interspersion of wetlands in the agricultural matrix poses a significant challenge for regional estimates of nitrous oxide (N2O) or methane (CH4) flux. For example, the meta-analysis of the carbon balance of North American wetlands published by Bridgham et al. (2006) contained few N2O and no CH4 estimates from the Prairie Pothole region and none for either gas from the Canadian portion of the region; no studies on CH4 for the Canadian portion have appeared since their analysis. Given the extent of the Prairie Pothole region (estimated at approximately 750,000 km2) it is essential that we refine our flux estimates.

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In wetlands, the absence of oxygen created by saturated conditions causes alternative electron acceptors to be utilized for the oxidation of organic matter. In the sequential reduction sequence, O2 depletion is followed by sequential utilization of nitrate, manganese, iron, sulphate, and finally CO2 as electron acceptors (Alewell et al., 2008). Prairie wetlands also typically have high dissolved organic carbon (DOC) concentrations (Waiser, 2006), and the abundance of these potential electron donors may overwhelm the sequential reduction sequence and lead to utilization of several electron acceptors simultaneously (Alewell et al., 2008). The production of the important greenhouse gases nitrous oxide (N2O) and methane (CH4) is associated with reduction of nitrate and CO2 respectively and hence wetlands are known to be landscape-scale hotspots for emissions of these gases. In the topographically complex landscapes of the Prairie Pothole region, several N2O–generating processes may be operating simultaneously. The highest N2O fluxes in cultivated depressions of the Prairie Pothole region have typically been attributed to denitrification under SOC-rich, oxygen-limited (anaerobic) conditions during spring snowmelt or following a major precipitation event (van Kessel et al., 1993; Corre et al., 1996, 1999). In a process-based study at the SDNWA, Bedard-Haughn et al. (2006a) noted that denitrification was the dominant N2O-emitting process (N75% of total N2O release) from riparian vegetation zones in uncultivated ephemeral wetlands, whereas 50% or more of the N2O emissions from cultivated ephemeral wetlands were derived from nitrification or nitrifier denitrification. These differences in N2O emissions from vegetation zones can be difficult to detect — Phillips and Beeri (2008) did not find differences in N2O emissions from wetland vegetation zones in the Prairie Pothole region in North Dakota due to the high variability in N2O emissions on the two days that they measured emissions on in 2003. Hence the influence of wetland vegetation cover on the processes responsible for N2O emissions and annual amount of emissions remains unresolved. Unlike N2O, there are no published annual estimates of CH4 emissions from freshwater mineral wetlands of the Prairie Pothole region (Bridgham et al., 2006; Phillips and Beeri, 2008). Our overall understanding of the controls on methane production by soils and wetlands continues to evolve. The flux of CH4 is primarily controlled by the balance between methane production by methanogenic bacteria, which occurs when oxygen has been depleted in the anaerobic soil layers, and methane consumption (oxidation) by methanotrophic bacteria, which occurs in aerobic soil layers (Conrad, 1996; Whalen, 2005). Increasingly, the importance of anaerobic methane consumption by syntrophic communities of anaerobic bacteria and archaea is also being recognized (Stams and Plugge, 2009). For wetlands the primary hydrological control on flux appears to be methane consumption in an aerobic layer in the soil or sediment, and hence methane flux is highly dependent on water table depth (Jungkunst and Fiedler, 2007). In their recent meta-analysis, these authors found that CH4 emissions increase greatly when water table is 10 cm or less from the surface. Hence in these wetlands the extent and duration of water inundation and the presence of alternative electron acceptors in water and soils are major hydrological controls on the production of N2O and CH4. The extent and duration of water saturation in the wetlands of the Prairie Pothole region are known to be very dynamic and substantial interannual and within-annual variation in water depths can be seen in the SDNWA record (Fig. 1). Our primary objective was to capture the contributions of a semipermanent and several ephemeral wetlands to N2O and CH4 emissions from an agricultural landscape of the Prairie Pothole region over a 3-year period. The GHG measurements are coupled with observations on pond hydrology and chemistry to examine the interactions between the GHG processes and major landscape controls — specifically the effect of the extent and duration of inundation in the wetlands, and of the concentration of an alternative electron acceptor, sulphate, on emissions.

309

Finally, the role of riparian vegetation on emissions had been noted in previous, short-term studies, and we assessed the importance of vegetation type on emissions in our multi-year observations. 2. Methods and materials 2.1. Field site and sampling design The research site is a portion of the SDNWA, 52° 12′ N latitude, 106° 5′W longitude, located approximately 40 km east of Saskatoon, Saskatchewan, Canada. The area is a hummocky, glacial till terrain with some glacio-lacustrine sediments (previously described in Yates et al., 2006). The site consists of a ring of episodically connected wetlands and a cultivated upland located 5 to 10 m above the surrounding fringe of wetlands (Fig. 2a). The soils found at the site ranged from thin, low organic matter Typic Calciborolls (USDA Soil Taxonomy) (Chernozemic Rego Dark Brown (Canadian System of Soil Classification)) on shoulders, to thicker, organic matter rich Typic Haploborolls (Chernozemic Orthic Dark Brown) in midslopes and periodically water saturated soils such as Albic Argiborolls and Argic Cryaquolls (Chernozemic Eluviated Dark Brown and Gleysolic Humic Luvic) in depressions. Haplic Calciborolls (Chernozemic Calcareous Dark Brown) soils were found on some shoulders and adjacent to depressions. Soil texture ranged from loam at shoulder positions to silt loam in the depressions (Yates et al., 2006). Pond 1 is the largest pond at the SDNWA (Fig. 2) and is a semipermanent wetland (Class IV) dominated by an open-water phase devoid of emergent vegetation (Cover Type 4) (using the classification system of Stewart and Kantrud, 1971). The ephemeral wetlands of the upland were cultivated prior to 1968 (Bedard-Haughn et al., 2006b). For our study these wetlands were further sub-divided into three landscape elements after preliminary surveys of soils and vegetation. This stratification allowed evaluation of the importance of the classification by vegetation for annual emissions. The basin center is a level area characterized by non-grasses such as Mentha arvensis L., Cirsium arvense (L.) Scop., and Urtica gracilis Ait. The riparian grass is a non-level fringe area supporting grasses such as Bromus inermis Leyss. The riparian trees are a partial fringe of mixed trees and shrubs such as Salix spp., Populus balsamifera L., and Populus tremuloides Michx. Bedard-Haughn et al. (2006b) reports an area of 0.6 ha for basin center, 1.0 ha for riparian grass and 1.5 ha for riparian tree elements for the 29.3 ha upland at SDNWA (or approximately 10% of the total area). Mean bulk densities in the wetland units ranged from 0.78 to 0.93 g cm− 3 but did not differ significantly among landscape elements. 2.2. Wetland sampling design In July of 2003 a 31-point stratified sampling design was established in the wetlands and on the adjacent cultivated land. Five wetlands were selected and short transects of 5 to 9 points were established on the long axis of each of these wetlands (Fig. 2c). The transects were placed to ensure multiple replicated measurements were taken from each of the three wetland elements zones (riparian tree, riparian grass, and basin center) across the five ponds. In each wetland, docks were constructed to ensure that basin center landscape elements could be sampled without disturbance of the sediment surface when the basin centers were inundated. The cultivated field surrounding the wetlands was fallow in 2002 and 2003 and in May of 2004 was seeded to grass by Ducks Unlimited Canada. The mix consisted of Agropyron elongatum (Host) Beauvois, Agropyron intermedium (Host) Beauvois, Bromus biebersteinii Roem. and Schult., Elymus dauricus Turcz. exgriseb., Festuca rubra L., Onobrychis viciifolia Scop., Elymus Canadensis L., Agropyron trachycaulum (Link) Malte and Medicago sativa L. The one exception was the area surrounding Pond 117 (beyond the riparian zone), which was under

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Fig. 1. Pond water levels from 1968 to 2006 for selected ponds at SDNWA: a) Pond 1 (semi-permanent/permanent), and three ephemeral ponds b) 105, c) 117, and d) 120. Data from G. Van der Kamp, Environment Canada).

conservation tillage and was seeded to barley (Hordeum vulgare L.) in 2003, canola (Brassica napus L.) in 2004, spring wheat (Triticum aestivum L.) in 2005, and flax (Linum usitatissimum) in 2006.

2.3. Gas sampling design Gas sampling at Pond 1 was carried out from 2003 to 2006 (Table 1). In late 2003 a 100-m dock that started at the water's edge was constructed and a transect of 12 chambers (six each side) was located along it. The acrylic, non-vented chambers had a headspace volume of 10.76 L covering a surface area of 0.06 m2. Each chamber was attached to a fixed rod adjacent to the dock with a saddle clamp, which allowed chambers to be lowered onto the water's surface for the sampling period and stored off of the surface. Two riparian transects of 12 PVC chambers (as described below) each were also installed. The riparian area was inundated in 2005 and 2006 and no riparian sampling was completed. A rapid rise in water level after snowmelt in 2005 (Fig. 1) destroyed the dock used in 2004, thus necessitating the implementation of a different sampling design. Six sampling stations were accessed from a wooden dock that started at the water's edge; ten additional chambers were accessed from two

Table 1 Sampling period and frequency of gas sampling at St. Denis National Wildlife Area. Pond 1 Ephemeral ponds Duration of sampling Period of sampling Number of sampling days Days Fig. 2. Site map of the St. Denis National Wildlife Area: a) aerial photograph showing Pond 1 and the upland research area inside highlighted rectangle (total area shown is approximately 1.4 km × 1.4 km), b) upland research area showing the main ponds sampled (area is approximately 800 m × 600 m), and c) pond labels for ephemeral ponds (area is approximately 800 m × 800 m).

2003 9 2004 28 2005 15 2006 10

9 21 25 31

108.5 232.5 231.5 205.4

3 July to 10 October 5 March to 31 October 15 March to 18 October 4 April to 11 October

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floating platforms located in the open water in the middle of Pond 1 (five chambers each). Two types of chambers were used for gas sampling in the riparian areas. If there was no standing water on the surface soil gas flux was measured using a two-piece, closed, vented chamber consisting of a polyvinyl chloride (PVC) ring base and vented cap with a sampling port (similar to Hutchinson and Mosier, 1981). The chamber, when placed for sample collection, had a head space of 2.25 L and covered a soil surface area of 0.02 m2. Prior to the start of data collection, the bases were pressed into the soil and secured using 20-cm spikes where they remained for the duration of the field season. Where there was standing water on the surface (including basin centers) measurements were taken from the surface of the water using the acrylic, non-vented chambers that were also used at Pond 1. Preliminary testing was done to ensure both chambers provided comparable flux readings (see also Matson et al., 2009). Samples of the headspace gas were drawn from each type of chamber with a 20-mL syringe and injected into 12-mL evacuated tubes for transport back to the laboratory. Samples of the headspace gas were drawn at three equally spaced time intervals of 8 min. Sampling was completed in less than 1 hour and was timed to occur between 1200 and 1300 hrs. The sampling program was designed to capture the period from the onset of snowmelt through to the first snow fall in the following fall (Table 1). The number of sampling days was increased in 2006 to more fully capture emissions associated with the individual ponds at SDNWA. 2.4. Gas measurement and flux calculations N2O and CH4 samples were analyzed using a Varian CP-3800 GC (Varian Canada Inc., Mississauga, ON) with dual electron capture detectors (ECD's) for N2O and a flame ionization detector (FID) for CH4. N2O separations were carried out by Poraplot Q fused silica columns (12.5 m length× 0.32 mm diameter, film thickness= 8 µm); operating conditions for the GC were: 70 °C injector temperature, 35 °C oven temperature and 370 °C detector temperature. CH4 separations used a Porapak Q8 column (3.7 m length × 0.3 cm diameter, film thickness = 2 mm); operating conditions for the GC were: 70 °C injector temperature, 50–200 °C oven temperature and 200 °C detector temperature. The carrier gas was ultra-high purity He (7.9 and 14.4 mL min− 1 for ECD's, 40 mL min− 1 for FID) with P5 (95:5 v/v Ar:CH4 mix) as the makeup gas for the ECD's (10 and 12.0 mL min− 1). Samples (N2O:300 µL, CH4:900 µL) were introduced using a CombiPAL auto-sampler (CTC Analytics AG, Switzerland) with on-column injection and a split ratio of 10:1 (carrier gas+ make-up gas: gas sample) for the ECD's. Data processing was performed using the Varian Star Chromatography Workstation (ver. 6.2) software. Internal calibration curves were acquired by applying linear least squares regression to the gas concentration versus peak area data; gas concentrations in the samples were then calculated from the regression equations. Ambient air samples (2 in every 50 gas samples) and gas standards (1 in 50) were included in each analytical run, and used to check precision, correct for detector drift and calculate the minimum detectable concentration difference (MDCD). The MDCD was calculated from the average of pairs of ambient samples using the following equation (Yates et al., 2006): MDCD = cpair

diff

+ ð2σpair

diff Þ:

ð1Þ

c = average difference between sample pairs σ = standard deviation between sample pairs When calculating the gas flux, the MDCD was used to determine if the samples for each time step were significantly different from the t0; if the concentration difference between the t0 and each time step

311

was bMDCD, they were not considered different. If they were different, a polynomial equation was fit to the curve that the concentration from the four time steps generated. The flux was calculated as the first derivative of the second-order polynomial equation (y = ax2 + bx + c) used to describe the relationship between concentration and time. The MDCD and flux calculation are described in detail by Yates et al. (2006). The cumulative annual emissions for each site were calculated by multiplying the flux rates by the time that elapsed between each measurement (thereby assuming that emissions were constant throughout the day and between sampling dates). This calculation cannot account for emissions that occurred outside the sampling season. 2.5. Measurement of soil moisture, soil temperature and climate data Snow surveys were completed each winter in early March. The depth of snow was measured at each gas sampling point and snow cores using a fixed volume corer were taken at every second point for the calculation of the water equivalent of the snow pack. Volumetric soil moisture was measured at each location within an hour of each gas sampling, over a 15-cm depth, using time domain reflectometry (Topp and Ferre, 2002) and read manually from a Tektronix 1502 B cable tester (Tektronix, Wilsonville, OR). Volumetric moisture and bulk density (from Bedard-Haughn et al., 2006a) were used to determine water-filled pore space (% WFPS). Soil temperature, at the 5-cm depth, was obtained using a buried type T thermocouple constructed of twisted copper and constantan wire pairs and read using a Barnant DuaLogR™ thermocouple reader (Barnant Company, Barrington, IL). Precipitation and air temperature were measured at a station close to Pond 118 and averaged hourly. 2.6. Water depth and chemistry Water depths were measured bi-weekly at the deepest point in each pond. At each sampling point on each sampling day, the presence or absence of standing surface water was recorded. The pond water in each pond was sampled at three times each year (mid-April, midJune, and mid-August) during the ice-free season (unless dry down of the pond occurred prior to the third sampling). Samples were analyzed for pH, conductivity, ammonium, nitrate, total phosphorus, and major anions and cations using standard techniques (APHA, 1998) by ALS Laboratory Group, Edmonton, Alberta. The water samples were also assessed for DOC concentrations. All samples were initially filtered with a 1.5 µm glass-fibre filter (Whatman 934-AH) followed by a 0.7 µm glass-fibre filter and then were refrigerated and transported to the National Water Research Institute for analysis. DOC concentrations were determined using an Apollo 9000 Total Carbon Analyzer (Techmar–Dohrmann) employing platinum on aluminum oxide catalyst at 900 °C. 3. Results 3.1. Precipitation and air temperature Precipitation during the study was at (2004) or well above (2005, 2006) the 1971–2000 average in Saskatoon (Environment Canada, 2008) (Table 2). The average monthly temperatures during the study were generally within 2 to 3 °C of the 30-year average. 3.2. Hydrology of wetlands Pond 1 has a distinct hydrological regime compared to the other ponds studied at SDNWA. Long-term records from Environment Canada show that Pond 1 had never completely dried up in the period since 1968, whereas the period of inundation for the upland ponds ranged greatly between years (Fig. 1). In the years of our study Pond 1 experienced a very major increase in water level in 2005 and

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Table 2 Precipitation and temperature observations at the St. Denis National Wildlife Area. Year

2003 2004 2005 2006 30-Year Averagec a b c

Snow water equivalenta (mm)

Precipitation

Mean monthly air temperature

April 1 to Oct. 31 (mm)

March

76 55 68 118 84.8

252 282 406 355 281

April

May

June

July

August

September

October

3.9 5.8 6.9 4.4

7.8 9.7 11.6 11.5

13.2 14.3 16.7 16.0

18.3 16.4 17.6 20.5 18.2

19.5 13.7 14.9 17.5 17.3

10.1 10.4 10.9 12.2 11.2

5.5 2.7 NMb 1.9 4.5

(°C) − 5.8 − 6.4 − 8.0 − 5.8

Measured in mid-March. Not measured. Climate data from Saskatoon SK (Environment Canada, 2008).

maintained this water through 2006 (Table 3). In 2006 Pond 1 spilled over into a series of ponds downslope from the spillover point at its north-eastern edge. Similar year-to-year trends in water depth occurred in the upland ponds (Table 3). All of these ponds were dry when the project began in June 2003. In 2004 these ponds held (on average) about 55% of their maximum water depth and — except for pond 120, which is the largest of the ephemeral ponds studied — were dry by mid-May. All ponds held progressively more water in 2005 and 2006 and held this water later into the season — in the case of Pond 120 water remained in the pond throughout the year in both 2005 and 2006.

Unlike Pond 1 the range in water chemistry between years and between ponds in the upland ponds was relatively low (Table 4). The 2+ dominant anion in these ponds was HCO− and K+ were the 3 , and Ca dominant cations. SO2− levels were very low in these ponds. 4 Nutrient and carbon properties were more intensively sampled in 2006 in the ponds (Table 5). All ponds had nitrate-N levels below 0.006 mg L− 1 (data not shown) and the range for DOC among ponds was very small. The water in Pond 105 had notably higher levels of ammonium-N than the other ponds although the standard deviation for 105 was high relative to the other ponds. 3.4. N2O emissions from wetlands

3.3. Water chemistry of wetlands Pond 1 also had very distinctive water chemistry relative to the upland ponds (Table 4). In the low water conditions of 2004 Pond 1 had E.C. higher than 4500 µS cm− 1 and had SO2− 4 as the dominant anion and Mg2+ as the dominant cation. The large volume of snow meltwater in 2005 and 2006 diluted the salinity and concentrations of all the anions 2− and cations compared with 2004 except for HCO− 3 ; however SO4 and 2+ Mg remained the major anion and cation respectively.

Cumulative annual N2O emissions from Pond 1 were low (maximum annual emission b approximately 300 g N2O–N ha−1 year− 1) in the three complete years of the study (Table 6). In 2003 and 2004 the lower water levels in Pond 1 caused exposure of sediment along the shoreline. Emissions from the shoreline in 2004 (Fig. 3b) spiked in mid-April and again in mid-May but mean emissions never exceeded 10 ng N2O–N m−2 s− 1. The emissions from open water in 2004 (Fig. 3a) and 2005 (Fig. 3c) fluctuated around 0, and although a

Table 3 Pond hydrological observations for 2004 to 2006 and cumulative annual N2O and CH4 emissions for 2006 for ponds at St. Denis National Wildlife Area. Pond

1 103 105 117 118 120 a b c

1968–2006

2004

Maximum water depth (cm)a

Maximum water depth (cm)

Date pond dry by

Maximum water depth (cm)

2005 Date pond dry by

Maximum water depth (cm)

2006 Date pond dry by

Annual Emissions g N2O–N ha− 1

Annual CH4 emissions g CH4 m− 2

359 98 94 113 91 134

114 52 45 58 62 70

NAb May 19 May 19 NMc May 19 Aug. 11

336 80 81 113 85 123

All year Aug. 16 Oct. 19 Aug. 12 Aug. 16 NA

359 95 94 75 85 134

NA Aug. Aug. Aug. Aug. NA

310 676 2690 228 943 259

3.3 21.0 100.9 12.1 15.1 40.3

25 25 15 15

Maximum water depth for the pond recorded for 1968 to 2006 from National Water Research Institute, Environment Canada. Not applicable: Pond held water all year. Not measured in 2004.

Table 4 Water chemistry (mean and standard deviation) for ponds at the St. Denis National Wildlife Area. Pond

Period

Electrical cond. µS cm

Pond 1 Pond 1 Pond 1 103 105 117 118 120

2004 2005 2006 2004–06 2004–06 2004–06 2004–06 2004–06

−1

4530 ± 846 1830 ± 205 1057 ± 46.2 261 ± 69.8 284 ± 98 338 ± 192 224 ± 29.8 280 ± 116

SO2− 4 mg L

−1

3136 ± 44 267 ± 59 437 ± 93 5.48 ± 7.1 5.54 ± 6.3 6.18 ± 7.5 5.48 ± 4.8 5.62 ± 4.6

HCO− 3 mg L

−1

176 ± 122 194 ± 67 203 ± 62 141 ± 19.4 138 ± 12.3 144 ± 33.4 130 ± 16.7 133 ± 41.2

Cl− mg L

Ca2+ −1

68 ± 20 15 ± 12 12 ± 1.0 18.9 ± 24.4 17.2 ± 20.3 20.8 ± 24.4 4.53 ± 1.7 18.1 ± 23.9

mg L

−1

341 ± 95 72 ± 20 93 ± 6.7 25.9 ± 3.6 21.1 ± 3.4 25.4 ± 6.9 20.9 ± 3.2 25.6 ± 10.

Mg2+ mg L

−1

523 ± 148 53 ± 12 94 ± 22 8.49 ± 0.8 8.52 ± 0.9 8.92 ± 2.0 8.54 ± 0.9 8.15 ± 1.6

Na+ mg L

K+ −1

148 ± 44 13 ± 3.6 27 ± 6.0 0.98 ± 0.04 0.81 ± 0.27 1.17 ± 0.2 0.87 ± 0.2 1.21 ± 0.4

mg L− 1 172 ± 51 36 ± 12.5 36 ± 4.7 39.3 ± 28.9 42.3 ± 22.3 42.7 ± 30.2 22.1 ± 3.2 33.3 ± 29.9

D. Pennock et al. / Geoderma 155 (2010) 308–319 Table 5 Means and standard deviation for water chemistry measurements taken at ponds, St. Denis National Wildlife Area for 2006. Pond

1 103 105 117 118 120

Total P

Dissolved organic carbon

Dissolved inorganic carbon

NH4–N

mg L− 1

mg L− 1

mg L− 1

mg L− 1

0.20 ± 0.08 1.92 ± 1.51 2.10 ± 1.30 0.77 ± 0.38 0.84 ± 0.44 0.71 ± 0.30

29.9 ± 0.50 26.6 ± 11.0 30.1 ± 8.21 24.0 ± 11.2 22.9 ± 8.70 22.1 ± 7.20

44.5 ± 10.3 26.2 ± 11.5 28.8 ± 11.0 26.9 ± 17.0 27.8 ± 16.8 22.6 ± 11.9

0.04 ± 0.02 0.09 ± 0.10 0.42 ± 0.67 0.03 ± 0.01 0.04 ± 0.03 0.04 ± 0.02

sustained period of emissions occurred throughout June in 2006 (Fig. 3d), mean emissions were at or below 15 ng N2O–N m−2 s− 1 throughout the episode. Annual emissions for the three wetland landscape elements of the upland showed distinct differences in the cumulative amount and

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temporal pattern of N2O. Overall emissions in 2005 (not shown) closely resembled those for 2006 for all elements. Riparian Tree elements showed consistently low emissions in all three years compared to the other wetland elements (Table 6) and no period of sustained emissions in any year. WFPS was 70% or less in 2004, and emissions overall were very low in that year from this element (Fig. 4c). A single episode of high emissions occurred in 2006 in association with a rapid decrease in water-filled pore space from around 80% to less than 60% (Fig. 5c). The riparian grass elements in 2004 had mean emissions on each sampling day close to 0 ng N2O–N m−2 s− 1 (Fig. 4b) and cumulative emissions of only 92 g N2O–N ha−1 year− 1. In 2005 cumulative emissions overall increased to 520 g N2O–N ha−1 year− 1 but no period of higher emission occurred. In 2006 emissions from these elements showed an episode of sustained higher emissions that began after July 4 and ended by August 15 (Fig. 5b). This episode was associated with a rapid decrease in mean WFPS from greater than 80 to less than 60% that occurred over the same period (Fig. 5b) and contributed the majority of the 994 g N2O–N ha−1 year− 1 of cumulative emissions measured in 2006 from this element.

Table 6 Cumulative annual N2O and CH4 flux over measurement period for wetland landscape elements and Pond 1 at the St. Denis National Wildlife Area. The values for wetland elements are averaged across the ephemeral ponds. 2003a

2004

Landscape element

g N2O–N ha

Riparian grass Riparian tree Basin center Pond 1: riparian Pond 1: open water

350 279 172 425 0.6

a

2005

2006

−1

92 295 979 315 237

2003a

2004

2005

2006

Not measured Not measured Not measured 0.04 0.01

12.3 13.6 138.6 Not present 2.01

5.3 5.3 86.7 Not present 3.33

g CH4 m− 2 520 484 2099 Not present − 138

994 575 1484 Not present 310

Not measured Not measured Not measured 0.11 0.19

Part-year measurements.

Fig. 3. Seasonal variation in mean daily nitrous oxide flux from a) the open water of Pond 1 in 2004, b) the riparian area of Pond 1 in 2004, c) the open water of Pond 1 in 2005, d) the open water area of Pond 1 in 2006 Error bars represent the standard error of the mean.

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Fig. 4. Rainfall, mean soil temperature at 5 cm, mean water-filled pore space (WFPS) for 0- to 15-cm depth, and mean (and standard error of the mean) daily N2O emissions from b) riparian grass, c) riparian tree, and d) basin center elements for 2004. Triangles for soil temperature indicate the first sampling date for which the mean soil temperature at 5 cm was at or exceeded the indicated temperature.

The basin center elements showed the greatest inter-annual variations of any of the elements at the SDNWA (Table 6). All of the ponds were dry by the time sampling commenced in 2003, and cumulative emissions for this part year were very low (Table 6). In 2004 two minor episodes (mid-May and early July) occurred that had maximum emissions less than 20 ng N2O–N m−2 s− 1. The prolonged decrease in WFPS from saturation through to approximately 50% by the end of the season did not trigger a period of sustained emissions (Fig. 4d). In 2005 a sustained period of high emissions began in early August and continued through September. This episode was associated with the loss of water from most of the ponds (Table 3) and a rapid decrease in WFPS from greater than 80% to around 60%. In 2006 a similar but shorter duration episode occurred after surface water in the ponds was lost, as reflected by the sudden decrease in WFPS in late July (Fig. 5d). The effect of the loss of surface water can be seen for the detailed results for Pond 105 in 2006 (Fig. 6). When all parts of the basin center were covered with water, pond 105 has two periods of emission, with mean emissions around 20 ng N2O–N m−2 s− 1. Once surface water is lost from the periphery of the basin center and soil is exposed in early August, a period of sustained higher emissions from these exposed soils began, which continued through to early September. Total N2O emissions from this period were substantially higher than for the inundated period. The increased intensity of sampling in 2006 highlighted the considerable differences in cumulative emissions among the ponds (Table 3). Emissions were low from Pond 1 and 120, both of which held standing water throughout the measurement period. Cumulative emissions

Fig. 5. Rainfall, mean soil temperature at 5 cm, mean water-filled pore space (WFPS) for 0- to 15-cm depth, and mean (and standard error of the mean) daily N2O emissions from b) riparian grass, c) riparian tree, and d) basin center elements for 2006. Triangles for soil temperature indicate the first sampling date for which the mean soil temperature at 5 cm was at or exceeded the indicated temperature.

range from a low of 228 g N2O–N ha−1 year− 1 (Pond 117) to 2886 g N2O–N ha−1 year− 1 (Pond 105). 3.5. CH4 emissions from wetlands As with N2O there was a major difference in both the annual amount and the temporal patterns of emissions between Pond 1 and the ephemeral ponds at the site (Table 6). CH4 emissions from Pond 1 were low or very low (compared to the ephemeral ponds) in all three full years of the study. The lowest annual emissions occured in 2004, and emissions from both exposed sediment and open water were low in this year (Fig. 7). Although emissions increased 100-fold in 2005 and 2006 from 2004, the cumulative emissions were low compared to the ephemeral elements at the site. CH4 emissions in 2005 were higher from all three ephemeral elements than in 2006 (Fig. 8). The emissions in 2005 were primarily associated with a single, sustained period of emissions that began in early June in the basin centers and in late July in the other two elements. In 2006 two or more periods of emissions occurred in all three elements (Fig. 9). The temporal patterns from the aggregated data across ponds disguise considerable within- and between-pond differences in emissions (Table 3). Although the range in annual CH4 emissions from the ephemeral ponds were not as great as for N2O there was a still a 8X difference between the lowest emissions (Pond 117) and the highest (Pond 105). The temporal pattern for Pond 105 in 2006 (Fig. 6) showed three main phases: an initial period of negligible emissions from open water until late June; a period with two episodes of high emissions, the

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one, indicating consistently greater emissions in daytime (Wilcoxon Signed Rank test, significance 0.04). Daytime emissions were on average 1.3× higher than nighttime in the ephemeral ponds and 2.4× higher in Pond 1 (Table 7). This would suggest that the annual estimates for CH4 emissions from this site based solely on daytime values are overestimated by between approximately 15 and 25%. 3.7. Relationships between water chemistry and N2O and CH4 emissions There were few differences between Pond 1 and the ephemeral ponds in the substrates involved in the production of N2O (NH+ 4 , DOC) (Table 5). All ponds had nitrate-N levels below 0.006 mg L− 1 (data not shown). The major difference in CH4 emission between Pond 1 and the ephemeral ponds may be associated with the suppression of methanogenesis due to the presence of the competing electron acceptor, SO2− 4 . The dilution of Pond 1 through the years of the study provides contrasting SO2− conditions, with the low SO2− endpoint 4 4 formed by the ephemeral ponds (Table 4). The correlation between annual CH4 emissions and mean SO2− is significant (Pearson cor4 relation for log-transformed variables = − 0.9, p = 0.001), although a considerable range in CH4 emissions occurs for the low SO2− 4 ephemeral ponds (Fig. 10). The differences in N2O and CH4 emissions among the ephemeral ponds were not related to any of the water biogeochemistry measures (Table 5). The highest emissions in 2006 were associated with Pond 105, which also had the highest levels of total P, DOC, and especially NH+ 4 . However, the second highest emitting pond, Pond 120, did not exhibit higher mean levels for these properties. 4. Discussion 4.1. Pond hydrology and chemistry Fig. 6. Pond 105 water levels in 2006 and mean (and standard error of the mean) for a) CH4 from exposed wetland soil, b) CH4 from open water, c) N2O from exposed soil, and d) N2O from open water.

latter of which occurs during the final drying down of the pond and has high emissions from the remaining open water points and lower emissions from recently exposed sediments; and a final period with no emissions or very slight uptake after drainage of the water from the surface. 3.6. Diurnal effects on emissions The annual emissions presented in previous sections are based on extrapolation of measured observations. The observations were made mid-day: from approximately 1000 to 1500 on the sampling days. This introduces a bias into the results because this is also typically the period of highest temperature and hence possibly the highest rate of any biologically mediated process. To assess the extent of bias, observations were taken at two hour intervals for 20 to 24 h on 11 days during the sampling campaign and were used to examine if there were consistent differences between daytime (0600 to 1800) and nighttime (2000 to 4000) emissions from the two types of wetlands (Table 7). For N2O, both the ephemeral ponds and Pond 1 had a wide range of daytime: nighttime ratios but the direction of the ratios were not consistent: on four dates the ratio was greater than 1 and on seven days less than 1. The difference between paired nighttime and daytime emissions was not significant (Wilcoxon Signed Rank test, significance 0.68) and no correction was applied to the annual estimates. For the five days on which diurnal patterns of CH4 were assessed the ratios of daytime: nighttime emissions were consistently greater than

Our study captured the end of a dry period that extended from 1997 to 2004 which had lead to low water levels in all the ponds (Fig. 1). Water levels increased greatly due to a major spring runoff event in 2005 (Fig. 1) that was followed by a second high runoff year in 2006. Field observations made at SDNWA since 2004 have confirmed that surface water spillover (Leibowitz and Vining, 2003; Cook and Hauer, 2007) is primarily responsible for the substantial inter-annual variation in water level observed at the site. In the spillover process, the maximum pond water volume is determined by pond geometry and the minimum elevation of the drainage divide separating it from adjacent ponds (Winter and LaBaugh, 2003; Cook and Hauer, 2007). When the volume of water delivered to the pond through runoff exceeds this maximum storage capacity, the pond spills over into adjacent ponds which may in turn spillover, creating a cascading effect in the system. In the Canadian Prairie Pothole region the major annual runoff event is spring melt of the snowpack (van der Kamp and Hayashi, 2009). The groundwater flow system does contribute discharge to some ponds in these landscapes, but the volume of water contributed by discharge is relatively small (van der Kamp and Hayashi, 2009). The spillover mechanism is the major controller of pond hydrology in landscapes like that at SDNWA but the pond solute load (and hence the presence of alternative electron acceptors such as sulphate) is strongly influenced by the small volume of groundwater discharge (van der Kamp and Hayashi, 2009). Previous studies have noted two major types of water chemistry in ponds of 2− this region: HCO− 3 - dominated and SO4 -dominated (Driver and Peden, 1977; Waiser, 2006). Pond 1 is a SO2− 4 -dominated pond, and the concentration of SO2− in each year was strongly influenced by 4 the dilution effect of the water volume increase observed in 2005 and 2006 (Table 4). The only significant source of SO2− in these 4 landscapes is through oxidation of pyrite from underlying glacial till

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Fig. 7. Seasonal variation in mean daily methane flux from a) the open water of Pond 1 in 2004, b) the riparian area of Pond 1 in 2004, c) the open water of Pond 1 in 2005, d) the open water area of Pond 1 in 2006 Error bars represent the standard error of the mean.

Fig. 8. Mean (and standard error of the mean) daily CH4 emissions from a) riparian grass, b) riparian tree, and c) basin center elements of ephemeral ponds for 2005.

Fig. 9. Mean (and standard error of the mean) daily CH4 emissions from a) riparian grass, b) riparian tree, and c) basin center elements of ephemeral ponds for 2006.

D. Pennock et al. / Geoderma 155 (2010) 308–319 Table 7 Mean nighttime (2000 to 0400) and daytime (0600 to 1800) flux of N2O and CH4 from ponds at SDNWA. Date

Number of points

Nighttime

Daytime

Daytime: Nighttime

N2O (ng N2O–N m− 2 s− 1) from ephemeral ponds April 6 2004 5 1.94 ± 2.38 April 28 2004 10 77.5 ± 15.41 May 8–9 2004 20 0.75 ± 3.37 July 28 2004 16 0.92 ± 0.55 May 18–19 2006 18 1.90 ± 0.53 June 28 2006 20 8.92 ± 18.53

10.40 ± 4.35 73.70 ± 11.41 2.67 ± 8.23 0.72 ± 0.78 −14.1± 5.33 −1.02± 6.48

5.36 0.95 3.58 0.78 − 0.74 − 0.11

N2O (ng N2O–N m− 2 s− 1) from Pond 1 May 31 2005 6 6.16 ± 29.09 July 5 2005 6 2.98 ± 35.34 August 11 2005 6 4.95 ± 25.54

−1.20± 14.90 1.90 ± 58.40 0.97 ± 27.20

0.19 0.64 0.20

CH4 (µg CH4 m− 2 s− 1) from ephemeral ponds May 18–19 2006 18 0.68 ± .035 June 28 2006 20 3.72 ± 1.45

0.80 ± 0.22 5.19 ± 3.99

1.18 1.40

CH4 (µg CH4 m− 2 s-1) from Pond 1 May 31 2005 6 0.01 ± 0.02 July 5 2005 6 0.26 ± 1.03 August 11 2005 6 0.06 ± 0.03

0.02 ± 0.04 0.38 ± 1.16 0.23 ± 0.88

1.83 1.47 3.84

(Van Stempvoort et al., 1994), and the SO2− dominated ponds are 4 associated with the slow discharge of SO2− 4 -rich groundwater (van der Kamp and Hayashi, 2009). In ponds that receive no groundwater discharge such as the ephemeral ponds in this study, water chemistry is primarily controlled by the precipitation chemistry and by solutes leached by lateral flow from the surrounding catchment. The ephemeral ponds were 2− dominated by HCO− 3 , and SO4 concentrations are very low (Table 4). 4.2. Annual emissions of CH4 The annual emissions of CH4 observed in our study for the ephemeral freshwater mineral wetlands are high, especially for the basin center elements (Table 6). Bridgham et al. (2006) report a geometric mean measured emission from freshwater mineral wetlands in North America of 8.1 g CH4 m−2 yr− 1 but their summary includes few if any wetlands from the Prairie Pothole region. Bridgham et al. (2006) averaged across sites when multiple sites of the same wetland type were presented in the original papers used in their metaanalysis. Using this same approach, the average annual CH4 emissions across the three elements at SDNWA were 54.8 g CH4 m−2 yr− 1 in 2005 and 32.4 g CH4 m−2 yr− 1 in 2006. Even if the annual emissions

Fig. 10. Scatter plot of relationship between log of mean SO2− 4 concentrations of SDNWA ponds and log of mean annual flux of CH4 from the ponds. Values shown are for Pond 1 in 2004, 2005, 2006 and for the five ephemeral ponds in 2006.

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are reduced by 20% to remove the bias introduced by mid-day sampling, the annual emissions are well above the mean emissions from North American freshwater mineral wetlands. The high annual CH4 emissions from the ephemeral wetlands also contrast strongly with the low emissions (relative to the Bridgham et al., 2006 meta-analysis) from Pond 1 in all years of the study (Table 6). The differences in CH4 emissions between Pond 1 and the ephemeral ponds are probably related to the concentration of SO2− 4 (Fig. 10), although the relationship needs to be extended to more sites before the correlation can be reliably established. The effect of SO2− 4 on the suppression of methane production has been amply documented and summarized (Segers, 1998).The reduction of alternative electron acceptors such as SO2− can reduce substrate concentrations 4 to a value which is too low for methanogenesis or alternatively cause a stabilization of redox potential at levels too high for methanogenesis. Methane oxidation may also be coupled to reduction of sulphate during anaerobic oxidation of methane (Stams and Plugge, 2009), which may explain, in part, the observed absence of significant CH4 emission early in the season from the ephemeral ponds. Heagle et al. (2007) found that for the ephemeral pond 109 at SDNWA in 1994, substantial reduction of SO2− 4 occurred between spring melt and early July, which −1 reduced SO2− to 4 levels in the pond water from approximately 3 mg L −1 2− 0 mg L The removal of SO4 through reduction may trigger the onset of significant CH4 production in these ephemeral ponds. The high CH4 production by methanogenesis is also facilitated by relatively high DOC levels in prairie wetlands (Waiser, 2006). The ranges of DOC observed in our study are comparable to those observed by Waiser (2006) for a range of wetlands in the study region and correspond to the values for the high DOC sites in the study by Alewell et al. (2008) in Germany. In the ephemeral wetlands such as those at SDNWA, the basin center elements are invaded by terrestrial vegetation during dry years (such as the 1998 to 2003 period) that then dies back due to flooding in wet years (such as 2004–2006) and may provide a continuous supply of C decomposition products during the period of inundation. The results for 2005 and 2006 (the two highest emission years) for the ephemeral ponds also must be placed in the context of the longer term hydrological record at the SDNWA. Overall the pond water levels during our study were at the highest levels observed in the 39 years of observation at SDNWA (Fig. 1). In 14 of those years ponds such as 105 held no water at any point (Fig. 1) and hence they may well have acted as slight CH4 sinks. Hence the years of observation of our study may constitute the optimum possible conditions for CH4 emission from these ephemeral ponds. 4.3. Annual N2O emissions There were too few N2O studies from freshwater mineral wetlands for Bridgham et al. (2006) to summarize but our results can be compared to results from agricultural landscapes in western Canada. For a large regional study in the Black soil zone of Saskatchewan (Corre et al., 1999) calculated area-weighted emissions of 950 g N2O–N ha−1 yr− 1 and Pennock and Corre (2001) report an area-weighted annual emission of 970 g N2O–N ha−1 yr− 1 for a site that had similar topography, soils, and climate to the SDNWA. Overall the annual N2O emissions for Pond 1 are well below those observed for agricultural landscapes in the Prairies whereas the values for the three elements for the ephemeral ponds are comparable to those observed elsewhere. N2O production through a denitrification pathway occurs at substantially higher Eh values than reduction of SO2− and hence 4 unlike CH4 N2O production should not be related to SO2− levels 4 (Conrad, 1996). A likely reason for the low observed N2O emissions from Pond 1 is the sequential reduction process that occurs in denitrification itself — in saturated conditions the reduction process is very likely to be taken to the N2 endpoint and little N2O is released (Veldkamp et al., 1998).

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The extent and duration of inundation are more significant factors for the N2O emissions from the ephemeral ponds. The major N2O emission events in these ponds are associated with the transition from complete or near-complete saturation of the soil (i.e., WFPS of 80 to 100%) to WFPS less than 60% (Figs. 5 and 6). At WFPS N90% the proportion of N2 produced by denitrification relative to N2O increases greatly (thereby limiting N2O release) and at WFPS around 80% the production of N2O by denitrification is greatest. As WFPS decreases to approximately 60%, denitrification declines and nitrification (autotrophic and heterotrophic) can contribute significantly to N2O emissions (although typically at lower emission rates than those due to denitrification) (Yates et al., 2006). 4.4. Effect of wetland vegetation type on emissions The final landscape control examined was the effect of riparian vegetation on emissions. The flooding of the riparian zone in Pond 1 in 2005 eliminated it from the measurements, although the results for 2003 and 2004 indicated the potential of higher N2O emissions from the riparian fringe. N2O emissions from all three wetland elements of the ephemeral ponds were low in 2003 but were substantially higher from basin center elements than the two riparian elements in the wetter years of 2004 to 2006 (Table 6). The higher emissions from basin center elements were also very evident for CH4 in 2005 and 2006 (Table 6). For the ephemeral ponds in all years, the emissions of both gases from the two types of riparian vegetation (tree and grass) were very similar. Hence although previous studies (Bedard-Haughn et al., 2006a; Phillips and Beeri, 2008) had indicated a potential difference in emissions linked to riparian wetland vegetation type this was not of major significance in the three-year record of our study. 5. Conclusions Overall the interactions between-pond hydrology, water chemistry, and the various GHG producing processes is a classic example of a landscape-scale problem, where the soil processes and properties cannot be understood in isolation from their temporal and spatial content (Pennock and Veldkamp, 2006). The lateral transfers of water, solutes, and sediments at present or in the past are central to understanding the processes responsible for GHG at this site. The most important landscape control on emissions appears to be the effect of hydrology on both the period of inundation (in the ephemeral ponds) and the presence and concentration of alternative electron acceptors such as SO2− (in all ponds). Significant methane 4 production occurs in the ephemeral ponds during the period of inundation, and the drawdown of water from the surface that follows the period of inundation is the period of highest N2O emissions. The significant sub-decadal fluctuations in pond water levels that occur at the site also effects the abundance of the organic substrates in the ponds, which drive the reduction processes linked to CH4 and N2O release. Hence any changes in the timing and amount of major hydrological events in this region is likely to have a very significant impact on emission of both gases from these landscapes. Pond 1, the only large semi-permanent wetland at the site, acted as a small source or sink of N2O and CH4 in each year of the study. Although unreplicated in our study, its chemistry is very consistent with other large semi-permanent or permanent ponds in the Prairie Pothole region and further studies will establish if this is a consistent trend for this class of wetland. The correlation between SO2− and CH4 could be a very 4 useful tool for regional extrapolation across a broad range of wetlands in this zone of the Prairie Pothole region and closely links CH4 to the landscape-scale surface and groundwater redistribution patterns, but further assessment of the relationship in other sites is required. The freshwater mineral soil wetlands in the Prairie Pothole region continue to undergo significant land use conversion, with restoration of cultivated wetlands for wildlife habitat occurring simultaneously with

conversion of native wetlands to agricultural production. The development of more regionally specific estimates for carbon gain through restoration and its effects on CH4 and N2O emission is essential for the effect of these conversions on regional GHG estimates to be clarified. Acknowledgements The research program at the SDNWA was funded by the Natural Sciences and Engineering Research Council Strategic Grants program, BIOCAP Canada, Ducks Unlimited Canada, and Environment Canada. The field program was initially supervised by Dr. J. Braidek, and major field assistance was provided by S. Corbett, K. Rudolph, A. Taylor, and T. Brannen. Gas analyses were completed by D. Richman and N. Webb. K. Elliott coordinated all aspects of data analysis. Dr. Bob Clark of the Canadian Wildlife Service of Environment Canada provided considerable logistical support for this project. Dr. Rick Bourbonniere and Dr. Garth Van der Kamp (both of the National Water Research Institute, Environment Canada) provided the DOC (Bourbonniere) and water level (Van der Kamp) data. References Alewell, C., Paul, S., Lischeid, G., Storck, F.R., 2008. Co-regulation of redox processes in freshwater wetlands as a function of organic matter availability? Science of the Total Environment 404, 335–342. APHA, 1998. Standard Methods for the Examination of Water and Waste Water, 20th ed. American Public Health Association, Washington, D.C. Bedard-Haughn, A.K., Matson, A.L., Pennock, D.J., 2006a. Land use effects on gross nitrogen mineralization, nitrification, and N2O emissions in ephemeral wetlands. Soil Biology and Biochemistry 38, 3398–3406. Bedard-Haughn, A., Jongbloed, F., Akkerman, J., Uijl, A., de Jong, E., Yates, T., Pennock, D., 2006b. The effects of management history on soil organic carbon stores in ephemeral wetlands of hummocky agricultural landscapes. Geoderma 135, 296–306. 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