- Additional Plants: native plants, ornamentals, trees
- Crop Production: agroforestry, forestry, nutrient cycling
- Education and Training: decision support system, demonstration
- Natural Resources/Environment: riparian buffers, riverbank protection
- Production Systems: agroecosystems, holistic management
- Soil Management: nutrient mineralization, organic matter, soil analysis, soil quality/health
We measured nitrous oxide (N2O) fluxes and nitrogen (N) input in riparian buffer and a crop field soils located in the Bear Creek watershed in central Iowa. The results indicate that N2O emissions from soils in the riparian buffers were significantly less than those in the crop field and the ratio of N2O emission to N inputs in the riparian buffer soils was smaller than those in the crop field soils. This study suggests that riparian buffers should not be considered a major source of N2O emission in the watershed.
Non-point source (NPS) pollutants such as sediment, N, phosphorus (P) and pesticides are major causes of water quality problems worldwide (Duda, 1993; Tonderski, 1996; Carpenter et al., 1998). Shortly after the Waikato Valley Authority in New Zealand (1973) first discussed the use of riparian buffers for the prevention of water pollution, a number of research projects were initiated to quantify the ability of riparian buffers to control NPS pollution (e.g. Lowrance et al., 1983; Peterjohn and Correll, 1984). Based on these and other studies, riparian buffers have been recommended as one of the most effective tools for coping with NPS pollution (e.g. Mitsch et al., 2001; Sabater et al., 2003; Hubbard et al., 2004).
Important functions of riparian buffers related to NPS pollution control are filtering and retaining sediment, and immobilizing, storing, and transforming chemical inputs from uplands (Schultz et al., 2000). Many studies have shown that riparian buffers can reduce eroded sediment delivery to surface waters by 70 to 95% (e.g. Lee et al., 2000, 2003), N fluxes by 5 to more than 90% (e.g. Kuusemets et al., 2001; Dukes et al., 2002) and P losses by 27 to 97% (e.g. Uusi-Kamppa et al., 2000; Kuusemets et al., 2001). Denitrification is recognized as the major mechanism for reducing NO3- within riparian systems, with removal generally ranging from 2–7 g N m-2 y-1 (e.g.; Groffman and Hanson, 1997; Watts and Seitzinger, 2000).
It recently has been hypothesized that increased denitrification within riparian areas may trade a water quality benefit for an atmospheric problem (Groffman et al., 1998) resulting from the greenhouse effect of N2O produced during nitrification and denitrification (Wang et al., 1976) and ozone depletion (Crutzen, 1970; Liu et al., 1977). The global warming potential of N2O is 298 times that of carbon dioxide (CO2) and 25 times that of methane (CH4) in a 100-year time horizon (Forster et al., 2007). Some studies (Groffman et al., 1998, 2000; Hefting et al., 2003, 2006) conclude that N transformation by riparian buffers with high NO3- loads results in a significant increase of greenhouse gas emission. Groffman et al. (2002) suggested that the Intergovernmental Panel on Climate Change (IPCC) inventory might be improved by including more measurements of riparian N2O fluxes.
Numerous studies have emphasized the role of vegetation in soil processes within riparian buffers. However, there are conflicting results regarding the relationship between vegetation type and denitrification rate in riparian buffers. While some studies (e.g. Hubbard and Lowrance, 1997; Verchot et al., 1997) have found higher groundwater nitrate removal or denitrification rates in forested riparian zones, other studies (Groffman et al., 1991; Schnabel et al., 1996) have found higher removal in grass dominated riparian sites. Still other studies (e.g. Hefting et al., 2003; Dhondt et al., 2004) have found no significant difference in groundwater nitrate removal or denitrification rate between forested and grass-dominated riparian sites. This variability suggests that there are questions about the relationship between vegetation types in riparian buffers and the emission of N2O from their soils and illustrates the need for additional studies in various regions of the country, in different landscape settings, and under different vegetation communities to quantify the emission of N2O from soils in riparian buffers (Walker et al., 2002).
Numerous studies have observed increased soil N2O emission following the wetting of dry soil in tropical areas (Nobre et al., 2001), semiarid areas (e.g. Wulf et al., 1999; Saetre and Stark, 2005), Mediterranean areas (Fierer and Schimel, 2002), dry tropical forests (García-Méndez et al., 1991; Davidson et al., 1993), savanna (Scholes et al., 1997), agricultural lands (e.g. Kusa et al., 2002; Mikha et al., 2005) and in laboratory studies (e.g. Appel, 1998; Hütsch et al., 1999). The increase rates ranged from 5-fold up to 1,000-fold (e.g. Prieme and Christensen, 2001; Saetre and Stark, 2005) and magnitudes of the episodic N2O emission increase varied depending on soil texture (Appel, 1998; Austin et al., 2004), soil water content (Appel, 1998), root responses (Cui and Caldwell, 1997), amount of added water (Ruser et al., 2006) and the characteristics and availability of substrates (e.g. Van Gestel et al., 1993; Schaeffer et al., 2003 ). Based on these studies, it is apparent that even a single wetting event could account for a large proportion of the annual emission of N2O (e.g. Prieme and Christensen, 2001; Nobre et al., 2001). Thawing frozen soils can also lead to increased N2O emissions (e.g. Herrmann and Witter, 2002; Müller et al., 2003). Although the duration of such elevated emission is limited mostly to a few days, they have been found to be an important source of the total annual emissions from agricultural land (e.g. Wagner-Riddle and Thurtell, 1998; Teepe et al., 2004), forests (e.g. Papen and Butterbach-Bahl, 1999; Teepe et al., 2000), and grasslands (Kammann et al., 1998). Matzner and Borken (2008) observed that the emissions of N2O after thawing frozen soils were in some cases significantly larger from arable soils than from forest soils. Such events usually occurred when soil temperature is near 0oC (e.g. Chen et al., 1995; Müller et al., 2003). Matzner and Borken (2008) stated that the increase in N2O emission after thawing increases with colder temperatures of frozen soil. In temperate regions, observed N2O emissions during freezing-thawing periods in spring may account for up to 70% of the total yearly N2O losses (e.g. Teepe et al., 2000; Regina et al., 2004). From these results, it is summarized that short-term N2O peak emissions following rewetting dry soils and thawing frozen soils contributes substantially to annual N2O emissions. Intergovernmental Panel on Climate Change Tier 1 methodology (2006) estimates soil N2O emission by multiplying N inputs by an emission factor in crop fields since N inputs are a source of N2O emission. However, the N input-based IPCC methodology for estimating N2O emissions may underestimate fluxes in the regions where frequent rewetting of dry soils and thawing of frozen soils occurs.
Therefore, studies assessing the contribution of peak emissions to annual N2O emissions and evaluating the current IPCC methodology are clearly needed to better understand annual N2O fluxes and the N cycle within these systems.
The objectives of this study were:
1) to compare N2O emissions from riparian buffer systems comprised of forest, warm-season grasses, and cool-season grasses and an adjacent crop field, and
2) to compare the measured N2O emissions with estimated ones using IPCC methodology.