Anthropogenic nitrogen (N) inputs have been found to influence emissions of greenhouse gases from a variety of ecosystems; however, the effects of N loading on greenhouse gas dynamics in lakes are not well documented. We measured concentrations of carbon dioxide (CO2), methane (CH4), and nitrous oxide (N2O) in 26 lakes in the Colorado Rocky Mountains (USA) receiving elevated (5 – 8 kg N ha−1 yr−1) or low (<2 kg N ha−1 y−1) levels of atmospheric N deposition. The mean CO2 concentration in surface waters was 27 μmol L−1 and did not differ between deposition regions. The CH4 concentration was greater in low-deposition lakes (167 nmol L−1) compared to high-deposition lakes (48 nmol L−1), while the opposite was true for N2O. The concentration of N2O in surface water averaged 29 nmol L−1 in high-deposition lakes compared to 22 nmol L−1 in low-deposition lakes. Nitrous oxide is of particular interest because it is more potent than CO2 as a greenhouse gas and because of its role in the destruction of stratospheric ozone. To understand the potential magnitude of lake N2O production related to atmospheric N deposition, we applied two published methodologies for determining emissions from aquatic ecosystems to available data sets. We estimated contemporary global N2O emissions from lakes to be 0.04 – 2 Tg N y−1, increasing to 0.1 – 3.4 Tg N y−1 in 2050. The contemporary estimates represent 13–95% of emissions from rivers and estuaries, suggesting that further research is required to better quantify emission rates from lentic ecosystems.
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 Lakes play an important role in the regulation of climate though the exchange of heat, water, carbon (C), and greenhouse gases with the surrounding terrestrial ecosystem and atmosphere [Williamson et al., 2009]. Both CO2 and CH4 fluxes result from microbial mineralization of organic matter imported from the watershed or produced in the lake. Nitrous oxide is produced as an intermediate product of denitrification, the microbial reduction of nitrate (NO3−) to N2 gas, and of nitrification, the microbial oxidation of ammonia to NO3− [Knowles, 1982; Wrage et al., 2001]. Per molecule, CH4 and N2O are 25 and 298 times more potent, respectively, than CO2 in terms of global warming potential for the 100-year time horizon [Forster et al., 2007]. Previous work has shown that production of greenhouse gases has been enhanced by N loading in a variety of ecosystems including forests, grasslands, agricultural fields and streams [Beaulieu et al., 2011; Liu and Greaver, 2009], but data for lentic ecosystems, such as lakes and reservoirs, are lacking.
 Elevated N nitrogen inputs may accelerate rates of C cycling and result in increased emissions of CO2 and CH4 from lakes [Tranvik et al., 2009]. Nitrogen loading may also increase microbial N2O production through enhanced nitrification and denitrification in soils and sediments [Seitzinger and Nixon, 1985; Ullah and Zinati, 2006]. Indeed, the majority of total N2O production by rivers and estuaries (0.3 – 2.1 Tg N y−1) is estimated to result from anthropogenic N loading [Kroeze et al., 2010]. The current atmospheric concentration of N2O is 319 ppb and it is increasing 0.3% per year due to human activities [Forster et al., 2007; Nevison et al., 2007]. Increases in N2O production are of particular concern because, in addition to being a potent greenhouse gas, N2O is currently considered the single most important ozone-depleting substance [Ravishankara et al., 2009].
 As fertilizer runs off from agricultural soils and is transported through groundwater, streams, and rivers to the oceans, a fraction of N leaves the ecosystems as N2O. Such N2O emissions are considered indirect, and under methodology of the Intergovernmental Panel on Climate Change (IPCC), are estimated using factors that are applied to anthropogenic N loading rates based on the ratio of N2O to NO3− in water [De Klein et al., 2006; Mosier et al., 1998]. Nitrous oxide emissions from lakes receiving agricultural runoff and other N sources are not included in greenhouse gas inventories. Further, global N2O emissions from lakes have not been estimated even though lakes cover the same surface area as rivers and have the potential for substantial N2O production due to their longer water residence times. The few studies that have investigated N2O dynamics in lakes have suggested that lakes are not significant sources of N2O [Huttunen et al., 2003; Mengis et al., 1997]. Consequently, the rate of N2O emission from lakes is poorly constrained.
 Here we report on concentrations of CO2, CH4, and N2O in alpine and subalpine lakes in the Colorado Rocky Mountains receiving elevated or low rates of atmospheric N deposition. We also measured production of these gases by sediments from the study lakes. Atmospheric deposition is an important source of N to oligotrophic high-elevation lakes that are otherwise not subject to human perturbations, such as land use change, wastewater, and direct runoff from agricultural fields [Burns, 2004]. We expected a positive relationship between the N deposition rate and concentrations of dissolved greenhouse gases and sediment gas fluxes because of increased microbial activity. Such increased microbial activity could include enhanced respiration, nitrification, denitrification, and methanogenesis. Heterotrophic microorganisms could be fueled by dissolved organic carbon (DOC) as an energy source because N deposition can increase concentrations of DOC in lake water through increased primary production in N-deficient catchments and lakes [Tranvik et al., 2009; Weyhenmeyer and Jeppesen, 2010].
2.1. Study Site
 We sampled 26 lakes located in the Rocky Mountains of Colorado, USA between June and August 2008 (Figure 1a). Selected lakes were within 6 km of a trailhead. The eastern slopes of the Rocky Mountains near the Niwot Ridge Long-term Ecological Research site (NWT) and Loch Vale Watershed Research site (LVW) receive atmospheric N deposition from fossil fuel combustion and agricultural sources [Burns, 2004; Nanus et al., 2003]. The rate of inorganic N (NO3− + ammonium) deposition has increased over the past 20 years to 5–8 kg ha−1 y−1 (Figure 1b; data from the National Atmospheric Deposition Program (NADP), http://nadp.sws.uiuc.edu/). Lakes in central and western Colorado near the Rocky Mountain Biological Laboratory (RMBL, Gothic, CO) and the Mountain Studies Institute (MSI, Silverton, CO) receive <2 kg ha−1 y−1 atmospheric inputs of N. Lakes near NWT and LVW are considered to be in the high-deposition region and lakes near RMBL and MSI are considered to be in the low-deposition region. The extent to which the sampled lakes receive N from groundwater or natural leaching is not known. With the exception of Lake Estes, catchments of the sampled lakes were unpopulated, so we assume there are no N inputs from sewage and agricultural runoff.
 Mean annual temperatures at high elevations are less than 2°C and lakes are generally ice-covered between November and June [Baron et al., 2000]. Englemann spruce and subalpine fir forests are found below tree line (∼3,300 m) and alpine tundra is found above tree line. The sampled lakes were small, generally ∼0.1 km2, except for Lake Estes. The watersheds occupy geologically diverse bedrock [Kent and Porter, 1980]. Precambrian-age granite, gneiss, and schist dominate the underlying geology in the vicinity of Rocky Mountain National Park. The geologic parent materials near RMBL include Mesozoic sedimentary rocks and Paleozoic metamorphics and intrusives. Extensive volcanic deposits and felsic gneisses and granites of the Uncompahgre formation characterize the San Juan Mountains surrounding MSI. Other studies performed simultaneously have found that atmospheric N deposition significantly influences concentrations of particulate and dissolved N and phytoplankton nutrient limitation in these lakes [Elser et al., 2009a, 2009b].
2.2. Field Sampling and Laboratory Procedures
 Thirteen lakes in each N deposition region were visited once during summer 2008 (Table 1). Fieldwork was staggered so that sampling of lakes in high- and low-deposition regions was not strongly skewed by date. High-deposition lakes were sampled in late June and late July 2008 and low-deposition lakes were sampled in early July and early August 2008. All lakes were sampled from an inflatable boat. The depth of each lake was measured using a portable echo sounder. Water samples were collected from just above the sediments using a battery-powered submersible pump fitted with tubing to take in water at 0.5–10 m. Water samples were filtered with Pall A/E glass fiber filters (Pall Corporation, Port Washington, NY, USA) and frozen until analysis for DOC on a Shimadzu TOC 5000. Nitrate plus nitrite (hereafter: NO3−) concentration was measured on a Lachat Quick Chem 8000 Autoanalyzer. Data on DOC and NO3− concentrations for a subset of lakes were previously reported [McCrackin and Elser, 2011].
Table 1. Study Lakes by N Deposition Level and Average Values for DOC and NO3−a
Lake Depth (m)
The results of a statistical test comparing these variables between deposition regions are shown.
 Surface sediments were collected using a LaMotte dredge from a water depth of approximately 10 m or at the maximum lake depth if <10 m. The dredge collected sediments from an area of 221 cm2 to a depth of ∼7 cm. We were not able to collect sediment from all lakes because rocks, debris, or macrophytes prevented the dredge from operating properly (Table 1). Sediments were returned to the laboratory and processed within 24 h of collection. For each lake, six analytical replicate 100 g subsamples of homogenized sediments were slurried with 80 mL of lake water collected from just above the sediments. Bottles were purged of oxygen with nitrogen gas (N2). To estimate denitrification rates, acetylene was added to half of the bottles (three per lake) to block the reduction of N2O to N2 [Yoshinari and Knowles, 1976]. After vigorous shaking, we collected 10 mL samples from the headspace volume (∼550 mL) at the beginning and end of 4 h incubations conducted at 4°C in the dark. Gas samples were analyzed for CO2, CH4, and N2O on a Varian CP-3800 gas chromatograph (Agilent Technologies, Santa Clara, CA, U.S.) equipped with an electron capture detector, a thermal conductivity detector, and a flame ionization detector [Crill et al., 1995]. For N2O we used CH4 and argon carrier gas at 30 mL min−1 with a pre-column (0.5 m × 1/8″ Hayesep N 80/100 ss) followed by a main column (2 m × 1/8″ Hayesep D 80/100 ss). For CO2 and CH4 we used N2 carrier gas at 30 mL min−1 with the same pre-column followed by a main column (2 m × 1/8″ Porapack QS 80/100 ss). Estimated ambient production of CO2, CH4, and N2O by sediments was determined as the accumulation of these gases for incubations that were not amended with acetylene [García-ruiz et al., 1998; Rudaz et al., 1999]. The denitrification rate was determined as the production of N2O during the incubations amended with acetylene and has been reported elsewhere [McCrackin and Elser, 2011]. Sediment gas fluxes are reported on the basis of dry sediment mass that was converted to an areal basis using the sediment bulk density for each lake [Richardson et al., 2004].
 For measurement of dissolved trace gasses in the water column, 750 mL glass serum bottles were filled using a battery-powered submersible pump fitted with tubing to take in water just below the surface to above the sediment at 10 m depth, or the maximum lake depth if <10 m. The temperature and dissolved oxygen content of the water at the sampling depth was measured with a YSI model 85 temperature-oxygen probe (YSI, Yellow Springs, Ohio, USA). At least one bottle volume was allowed to overflow prior to introducing a 60 mL headspace of ambient air. Bottles were sealed with a screw cap fitted with gray-butyl stopper and shaken vigorously for 1 min [Cole and Caraco, 1998; Mengis et al., 1997]. Immediately after shaking, a 20 mL gas sample was collected from the headspace with a syringe and injected into a 10 mL serum vial. Three replicates were collected at each sampling depth per lake. Prior to the injection of the environmental sample, the serum vials were flushed with helium, sealed with gray butyl stopper and aluminum crimp top, and evacuated. Gas samples were analyzed for CO2, CH4, and N2O on the Varian CP-3800 gas chromatograph as previously discussed.
 The percent saturation of CO2, CH4, and N2O in water samples was calculated as % saturation = ([gas]measured/ [gas]saturated) * 100, where [gas]measured is the measured concentration of each gas adjusted for the introduction of trace gases in the ambient air headspace during equilibration and where [gas]saturated is the saturated concentration of each gas reflecting the atmospheric concentration of each gas and the solubility constant of each gas at the measured water temperature [Weiss, 1974; Weiss and Price, 1980; Wiesenburg and Guinasso, 1979].
 Sediment water content was determined as mass loss after drying subsamples at 105°C for 48 h and organic matter (OM) content was determined as mass loss on ignition at 550°C for 4 h. Total C and N content of dried sediments were measured with a PerkinElmer CHN elemental analyzer (PerkinElmer, Inc., Waltham, MA, USA). Total phosphorus (P) content of combusted sediment was measured colorimetrically following extraction with 0.5 M hydrochloric acid using the acid molybdate technique [Lukkari et al., 2007]. Results of sediment analyses have been reported separately and are not included here in detail [McCrackin and Elser, 2011].
2.3. Statistical Analysis
 We performed t tests to compare sediment CO2, CH4, and N2O fluxes, concentrations of dissolved NO3−, DOC, CO2, CH4, and N2O between high- and low-deposition regions. We used the Wilcoxon signed-rank test to compare concentrations of dissolved CO2, CH4, and N2O between the lake surface and above the sediment. Multiple-linear regression was used to identify relationships among predictor variables and gas flux rates from slurry incubations and surface water concentrations of dissolved gases. Predictor variables for sediment gas fluxes included concentrations of NO3− and DOC, sediment OM, sediment C, N, and P content, and ratios of sediment C:N, C:P, and N:P. Predictor variables for dissolved CO2, CH4, and N2O in surface water were DOC and NO3− concentrations, sediment CO2 fluxes (for dissolved CO3), sediment CH4 (for dissolved CH4), and rates of denitrification and sediment N2O flux (for dissolved N2O). Models were selected by considering all subsets on the basis of adjusted R2 and goodness-of-fit using the Akaike information criteria. We evaluated multicolinearity and selected final models for which tolerance values were > 0.5 for all predictor variables. When necessary, response and predictor variables were transformed to improve normality. All statistical tests were performed using JMP (SAS Institute, Inc., Version 8.0.1) with α = 0.05.
3.1. Greenhouse Gases in Lake Water
 The mean concentrations of NO3− and dissolved N2O were significantly greater in high-deposition lakes relative to low-deposition lakes (Tables 1 and 2). In contrast, average concentrations of DOC, CO2, and CH4 were greater in low deposition lakes than high-deposition lakes. The surface water concentration of CO2 was positively related to concentrations of DOC and NO3− (Table 3). The methane concentrations was positively related to DOC but negatively related to NO3−. The concentration of N2O was positively related to NO3− (Figure 2). There was no correlation between sampling dates and measured concentrations of DOC and dissolved trace gases (R2 < 0.04, P > 0.3) and minimal correlation for NO3− (R2 = 0.15, P = 0.05). Across all lakes, mean concentrations of dissolved CO2 and CH4 were greater in water just above the sediment compared to surface water (P < 0.001), while mean concentrations of dissolved N2O did not differ between depths (P > 0.05). The surface waters of lakes were supersaturated with CO2, CH4, and N2O relative to the atmosphere. For CO2, the degree of saturation did not differ between deposition regions (P > 0.05). For all lakes, the mean CO2 saturation was 190% (SE ± 14%). The mean saturation of CH4 in surface water was 1,930% and 8,280% in high- and low-deposition lakes, respectively, a significant difference. The mean saturation of N2O was greater in high-deposition lakes compared to low-deposition lakes (156% and 138%, respectively). Lakes in the low-deposition region were generally thermally stratified whereas lakes in the high-deposition region were not. All lakes were oxic at the depth where sediments were collected.
Table 2. Average (and Standard Error) Concentrations and Saturation Percent of Greenhouse Gases in Surface Waters of the Study Lakesa
The results of a statistical test comparing these variables between deposition regions are shown.
 Sediment characteristics were reported previously in detail [McCrackin and Elser, 2011]. Briefly, there were no significant differences in sediment OM, total C, N, and P content, or ratios of C:N, C:P, and N:P between lakes in high- and low-deposition regions (P > 0.05). Sediment C, N, and P contents averaged 8.2 mmol g−1, 0.6 mmol g−1, and 0.1 mmol g−1, respectively for all lakes.
 Sediment production of CO2 under anoxic conditions averaged 0.16 mmol C m−2 d−1 and 0.32 mmol C m−2 d−1 for high- and low-deposition lakes, respectively, a significant difference (Table 4). Methane fluxes were greater in sediments from low deposition lakes, averaging 0.8 μmol C m−2 h−1, compared to high-deposition lakes, which averaged 2.2 μmol C m−2 h−1. Across all lakes, CO2 and CH4 sediment fluxes were negatively related to NO3− concentrations (Table 3). Nitrous oxide production in sediments was only observed for four of the sampled lakes and averaged 2.3 (±1.6 SE) μmol N m−2 h−1 for these lakes. The sediment N2O flux did not differ between deposition regions and was not predicted by any of the identified variables.
Table 4. Average (and Standard Error) Sediment Greenhouse Gas Fluxes for the Study Lakes and Results of Statistical Test Comparing Deposition Regions
 Rates of atmospheric deposition are 5 – 8 kg N ha−1 y−1 in the Colorado Front Range, which represents 25 – 40% of estimated N mineralization rates in high-elevation soils [Fisk and Schmidt, 1995]. In the eastern U.S. and Europe, deposition rates are as high as 11 – 20 kg N ha−1 y−1 [Bergström and Jansson, 2006; Tørseth and Semb, 1998] but represent only 5 – 17% of estimated soil N mineralization rates [Groffman et al., 2006; Zak et al., 2006]. While deposition rates in Colorado are less than those of other regions, they constitute a relatively large fraction of N that is cycled. As a result, ecosystems in alpine and subalpine regions of Colorado may be more sensitive to ecological effects of atmospheric N. Indeed, soil N content and soil N mineralization rates have increased in old-growth forests and the C:N ratio of foliage has decreased compared to reference sites [Baron et al., 2000]. In Colorado lakes, changes in the composition and biomass of the diatom community, altered stoichiometric ratios of N to P in the water column, and shifted phytoplankton nutrient limitation have been associated with N deposition [Elser et al., 2009a]. In the present study, we found evidence that concentrations of dissolved N2O were greater in surface waters of lakes subject to elevated N deposition, while the reverse was true for dissolved CH4. Methane production by sediments and surface water CH4 concentrations were negatively related to NO3−, while N2O concentrations were positively related to NO3−. Our data suggest that atmospheric N loading affects greenhouse gas dynamics in lakes by increasing NO3− concentrations.
4.1. Carbon Mineralization in Sediments
 Contrary to our expectations, we found no evidence that atmospheric N deposition has enhanced C cycling rates in lake sediments. Rather, sediment CO2 production was greater in low-deposition lakes compared to high-deposition lakes. For all sampled lakes, C released by sediment CO2 production was two orders of magnitude greater than that released through CH4 production. Further, there was no difference in sediment OM and C content of bulk sediments between regions, but the concentration of DOC was greater in low deposition lakes compared to high-deposition lakes. The higher concentration of DOC in low-deposition lakes is consistent with greater sediment C mineralization rates here compared to the high-deposition region. However, we expected the opposite result because of the fertilization effect that N deposition could have on primary production in the catchments and water column. Indeed, previous fieldwork at many of the same lakes found that concentrations of chlorophyll and seston C were significantly greater in high-deposition lakes [Elser et al., 2009b]. Such differences between studies could reflect seasonal or year-to-year differences in catchment or lake productivity. It is also possible that growth of heterotrophic bacterioplankton in low deposition lakes are N limited, allowing for DOC to accumulate in the water column [Taylor and Townsend, 2010]. Last, at current rates of atmospheric N loading, catchment-specific properties for the sampled lakes might have a greater influence on lake DOC concentrations than does N deposition, as has been observed in other regions [Hessen et al., 2009].
 Methane may represent 20 – 60% of total C mineralization in lake sediments [Bastviken et al., 2008]. In our sediment slurry incubations, however, CH4 production represented only about 7% of measured C mineralization. In fact, CH4 production by sediments for the sampled lakes was two orders of magnitude less than that reported for a eutrophic lake and at the low end of that reported for boreal lakes [Algesten et al., 2005; Bastviken et al., 2008; Liikanen et al., 2002]. The low rate of CH4 production we observed in sediment could be partially explained by temperature, as we conducted our incubations at 4°C, which is colder than temperatures for other studies. Methane production in peat soils is strongly sensitive to temperature, with Q10 values of 5.3–16 [Dunfield et al., 1993] and similar results could occur in sediments. It is also possible that our slurry incubations were not completely anoxic. During denitrification, ratios of CO2:N2 production are 1.5–6 [Groffman et al., 2006], but for our incubations the average ratio of CO2:N2O production (as a proxy for CO2:N2) was significantly greater at 95 (±26 SE). This suggests that there was available oxygen for microbial respiration. Methanogenesis is sensitive to reduction-oxidation potential and the presence of strong oxidants, such as oxygen or NO3−, will suppress CH4 production [Le Mer and Roger, 2001]. We also observed a negative relationship between water NO3− concentrations and CH4 production (Table 3). This might explain why CH4 production was greater in low-deposition lakes compared to high-deposition lakes. Further, published data for CH4 production in sediments are often collected from eutrophic lakes and lakes surrounded by peatlands or bogs, which would likely have anoxic conditions in the sediments and, thus, be more favorable for CH4 production [Huttunen et al., 2003]. Last, the generally low CH4 fluxes we observed could also be result from relative poor quality OM in sediments of the sampled lakes.
4.2. Greenhouse Gas Emissions From Lakes
 Lake gas emissions not only depend on the concentration gradient between the lake and the atmosphere, but also on the gas exchange coefficient [Wanninkhof et al., 1987]. We did not measure gas exchange rates for the sampled lakes, so we cannot determine gas emissions. However, the surface waters were supersaturated with CO2, CH4, and N2O, suggesting that the lakes are net sources of these gases to the atmosphere. The surface water concentration of CO2 was three orders of magnitude greater than that of CH4 or N2O; thus, CO2 is likely the dominant greenhouse gas emitted from the sampled lakes. Also, while CO2 is 25 and 298 times less potent than CH4 and N2O, respectively, in terms of radiative forcing, we estimate that CO2 would represent about 80% of the global warming potential of total greenhouse gases emitted from the sampled lakes [Forster et al., 2007; Liikanen et al., 2002]. The dynamics of CO2 in the surface water do not appear to be influenced by atmospheric N inputs. We did observe that surface water concentrations of CO2 were positively related to DOC and NO3− (Table 3), although there was no significant relationship with either predictor variable when considered individually. This is interesting because we found NO3− to be negatively related to sediment CO2 production and because other studies have found strong positive relationships between CO2 concentrations and DOC [Sobek et al., 2003]. It is possible that respiration by bacterioplankton and sediment bacteria respond differently to N loading, thus, it would be useful to further investigate the extent to which bacteria in sediments and pelagic areas contribute to CO2 in surface water of the sampled lakes. Overall, however, our data are consistent with the well-documented finding that most lakes are heterotrophic and sources of CO2 to the atmosphere [Cole et al., 1994; Tranvik et al., 2009].
 There are two major pathways for CH4 emissions from lakes, diffusion and ebullition (bubble flux) of gas produced in the sediments. Ebullition may account for 20–70% of CH4 emissions from sediments, especially in shallow, eutrophic lakes [Bastviken et al., 2008; Juutinen et al., 2009]. Our approach for measuring CH4 concentrations only reflects diffusive sediment fluxes and is at the very low end of the range reported for lakes using similar methods [Bastviken et al., 2004]. While there is a general lack of CH4 data for high-elevation lakes, Smith and Lewis  sampled five lakes in the Colorado Front Range, including one of the lakes we visited (Long Lake). The dissolved CH4 concentration we observed for Long Lake is within the range they reported. Overall, evidence presented here suggests that N deposition has reduced sediment CH4 production and concentrations of dissolved CH4. Methanogenesis does not appear to be a significant source of greenhouse gas in the sampled lakes, independent of atmospheric N inputs. However, methanogenesis may be a significant C mineralization pathway if methane oxidation is important. Indeed, up to 80% of CH4 produced in sediments may be consumed in the water column [Bastviken et al., 2008].
 Our study lakes were generally supersaturated with N2O and concentrations were comparable to other lakes, although the number of data sets from other lakes is limited (Table 5). The data presented here suggest that atmospheric N deposition has increased concentrations of dissolved N2O and potentially emissions of N2O from lakes, consistent with studies of other aquatic ecosystems [Beaulieu et al., 2011; Liikanen et al., 2003; Seitzinger et al., 1984]. It is important to note that studies of N2O dynamics in aquatic ecosystems have largely been conducted in the Northern Hemisphere in temperate and boreal climates, and that data for areas such as the tropics are lacking. The lakes we sampled are located in the same region with similar elevations and climate. Thus, while they are comparable to those reported in the literature, they may not be representative of other regions or land uses, such as those in urban or agricultural areas. Our understanding of the factors influencing N2O dynamics, as well as spatial and temporal patterns of N2O emissions would greatly benefit from further experimental research in aquatic ecosystems.
Table 5. Comparison of Dissolved N2O and Sediment N2O Fluxes Among Different Studies
 In lakes, N2O is produced by denitrification in sediments and by nitrification in sediments and the water column [Mengis et al., 1997; Wrage et al., 2001]. For the lakes we sampled, the microbial source of N2O in the sampled lakes is unclear. We expected a correlation between sediment N2O production and lake water concentrations of N2O. Denitrification-related N2O production, however, was only observed in sediments of four lakes. We do not know why there were no N2O fluxes for the majority of sediments we sampled. Similar assays conducted with sediments from lakes in Norway found significantly greater N2O fluxes in lakes that receive elevated levels of N deposition [McCrackin and Elser, 2010]. Incubation temperatures (4°C for Colorado and 15°C for Norway) could explain the differences between studies. If our assays were not completely anoxic, rates denitrification and related N2O production could have been repressed. Also, differences in N2O production may also be the result of incomplete denitrification in Norwegian lakes due to relatively higher NO3− concentrations. The sediments are still a possible source of N2O production because experiments conducted simultaneously found considerable denitrification capacity under NO3− enrichment [McCrackin and Elser, 2011]. Nitrification could produce N2O as has been observed in soils [Bateman and Baggs, 2005]. Assays conducted with sediments from Norwegian lakes in high-deposition regions showed significantly greater rates of nitrification potential (in response to non-limiting quantities of ammonium) than those in low-deposition regions (M. McCrackin, unpublished data, 2010). Also, nitrification in the water column is a potential significant source of N2O [Mengis et al., 1997; Twining et al., 2007]. Further investigation is required to determine the microbial sources of dissolved N2O observed in lakes. It is difficult to make generalizations of our data based on summer measurements, because we do not know how concentrations of dissolved CO2, CH4, and N2O in the sampled lakes vary seasonally. The alpine and subalpine lakes that we sampled are ice-covered for more than half of the year. Stream water NO3− concentrations in LVW of the Colorado Front Range tend to peak in lake May, decrease during June and July, and increase in the fall [Baron and Campbell, 1997]. Assuming lakes follow a similar pattern, our sampling likely missed peak NO3− concentrations. Because NO3− was an important predictor of gas concentrations and sediment gas fluxes, we might expect these variables to change seasonally. Gases that accumulate under ice during winter could be released during periods of thaw and mixing. Hence, concentrations of dissolved gases that we measured in summer months may be greater than average annual concentrations.
 While CH4 and N2O are more potent than CO2 in terms of radiative forcing, the low concentrations of these trace gases relative to CO2 indicate that lake greenhouse gas dynamics are dominated by CO2. The relatively small surface area of the sampled lakes suggests that they do not contribute disproportionately to such dynamics in the Colorado Rocky Mountains. Given that N2O plays a significant role in the depletion of stratospheric ozone, however, potential N2O emissions from lakes, and the effects of N deposition on them, deserve further consideration.
 The IPCC Guidelines for National Greenhouse Gas Inventories recognize that as N fertilizer leaches from agricultural soils and moves through groundwater and river networks, a fraction is converted to N2O. Such fluxes are considered indirect N2O emissions from N used in agriculture. Indirect emissions are estimated for groundwater, rivers, and estuaries using an emission factor (EF5) of 0.0075 kg N2O-N per kg N input, with an uncertainty range of 0.0005–0.025 kg N2O-N kg−1 N. The EF5 was originally based on the ratios of dissolved N2O-N to NO3− from groundwater and agricultural drainage water in published studies [Beaulieu et al., 2008; Mosier et al., 1998]. While lakes are not currently included in greenhouse gas inventories, we used the general approach for estimating emission factors as the ratio of N2O-N:NO3−. This ratio averaged 0.01 with a range of 0.001–0.07 for the lakes we sampled, somewhat larger than the default EF5 value of 0.0075 for aquatic ecosystems. Clarifying the relationship between N2O emissions and N loading rates is important, but the dynamics of N2O in lakes are not well documented. We surveyed the scientific literature and, where concentrations of N2O and NO3− were reported, determined the N2O-N:NO3− ratio to be between 0 and 0.12 (Table 6). This broad range indicates that a single N2O-N:NO3− ratio cannot be generalized across all lakes. Further, it is not known how the ratio of N2O-N:NO3− varies seasonally. For example, in regions where lakes develop ice cover during winter, allowing N2O to accumulate, ratios of N2O-N:NO3− at spring melt could be substantially greater than those in other seasons. Nonetheless, values for the high end of the range of N2O-N:NO3− ratios suggest there is potential for N2O emissions from lakes subject to elevated N loading. Indeed, Wang et al.  reported N2O production of 300 μmol N m−2 d−1 in pelagic areas of a hyper-eutrophic lake, which is an order of magnitude greater than maximum emissions reported for agricultural fields [Bouwman et al., 2002].
Table 6. Comparison of the Ratio of N2O-N:NO3− Among Different Studies
4.3. Global N2O Emissions From Lakes Receiving N Deposition
 Our data suggest a possibility that lakes could be an underappreciated component of global N2O cycling, especially given elevated N inputs. Here we attempt to quantify this possibility. Globally, the largest source of anthropogenic N2O emissions is the enhanced conversion of N fertilizer by microorganisms in agricultural soils [Forster et al., 2007]. Emissions from aquatic ecosystems, however, are also significant. Globally, rivers may convert 0.68 Tg N y−1 of anthropogenic N to N2O [Beaulieu et al., 2011]. Kroeze et al.  estimated contemporary natural and anthropogenic N2O emissions by rivers and estuaries to be between 0.3 and 2.1 Tg N y−1, which represent approximately 2–12% of total anthropogenic N2O emissions (∼17 Tg N y−1) [Denman et al., 2007].
 Lakes may receive N from a variety of sources other than atmospheric deposition, such as groundwater, sewage, agricultural run-off, and natural sources. All of these sources may contribute to lake N2O emissions. Here, we will estimate global N2O emissions related to direct N deposition to the lake surface and N deposition that is leached from the surrounding catchment. First, we first estimated global N loading to lakes via atmospheric N deposition using published data sets of N deposition rates (for 1993 and 2050) and of lakes and reservoirs with surface area >1 km2 [Dentener, 2006; Lehner and Döll, 2004]. We quantified N deposited directly to the lake surface (N kg y−1) = atmospheric deposition kg N ha−1 y−1 * lake surface area (ha). Estimated N delivered directly to lake surfaces is 1.1 and 2.1 Tg N y−1 for 1993 (contemporary) and 2050 (modeled), respectively. Based on our analysis, over 90% of the global surface area of lakes (2.6 million km2) is subject to natural and/or anthropogenic atmospheric N deposition. We accounted for watershed inputs of N deposition as (N kg y−1) = N deposition rate kg ha−1 y−1 * watershed surface area (ha) * 30%, where we assumed that that 30% of all N deposited to the watershed is subsequently leached to the lake [Intergovernmental Panel on Climate Change (IPCC), 2006]. Watershed area data were only available for large lakes (>50 km2). Based on the relationship of watershed surface area to lake surface area for large lakes reported by Lehner and Döll , we assumed for small lakes that the watershed area was 12 times greater than the lake surface area. Inputs of atmospheric N from watershed to lakes were estimated as 13.7 Tg N y−1 for 1993 and 26.1 Tg N y−1 for 2050. The extent to which lakes are subject to N deposition is of interest not only because of the potential for increased N2O emissions but also because of the documented effects of N on lake stoichiometry and food web dynamics [Elser et al., 2009b; Hessen et al., 2009]. Even lakes that are not directly influenced by human activity are at risk from atmospherically delivered pollution, which is of particular concern because N deposition rates are expected to increase globally in the next few decades, driven by energy demands and agricultural activities [Dentener, 2006].
 We next estimated N2O production using the method of Kroeze et al. , which assumes that emissions from rivers and estuaries result from nitrification and denitrification of inorganic N inputs as, N2O-N kg y−1 = (Nitrification + Denitrification) * EF. Here, anthropogenic and natural inorganic N sources are not distinguished. Seitzinger and Kroeze  describe this approach in detail. Briefly, it is assumed that 50% of N inputs are denitrified and that the nitrification rate exceeds the denitrification rate by 20%. The emission factor (EF) is 0.3% of total denitrification and nitrification except where N loading rates exceed 10 kg N ha−1 y−1, when the EF is 3%. Using this method we estimated lake N2O emissions related to total atmospheric N deposition to be 0.33 Tg N y−1 in 1993 and 0.9 Tg N y−1 in 2050.
 We also estimated N2O emissions using IPCC methodology for indirect greenhouse emissions from aquatic ecosystems. This approach is intended to estimate anthropogenic N2O emissions resulting from N run-off from agricultural systems and does specifically address atmospheric deposition to lakes. In our opinion, however, this methodology is most appropriate under current IPCC guidelines. Here, N2O emissions are determined as N2O-N kg y−1 = N inputs to lake kg N y−1 * EF5. Nitrogen inputs to lake are defined as atmospheric N deposition, calculated as previously described, and EF5 is 0.0075 kg N2O-N per kg N input (uncertainty range of 0.0005–0.025 kg N2O-N per kg N input). The resulting estimate for 1993 is 0.6 Tg N y−1 (range 0.04–2 Tg N y−1). Emissions from Asia and North America contribute about 44% and 23% of total global emissions, respectively, because of the relative high densities of lakes and regional patterns of N deposition (Figure 3a). Estimated 2050 emissions increase to 1 Tg N y−1 (range 0.07–3.4 Tg N y−1), with nearly half of this increase occurring in Asia (Figure 3b). Our 1993 estimates represent 13–95% of N2O emitted from rivers and estuaries reported by Kroeze et al. , although our estimates are not strictly comparable because we do not consider all sources of N inputs to lakes.
 Lake emissions calculated using the approach of Kroeze et al.  fall within the ranges determined using the IPCC methodology. It should be noted that the method used by Kroeze et al. includes all natural and anthropogenic sources of N2O, while the IPCC methodology is intended to quantify only anthropogenic emissions. Compared to modeled pre-industrial rates of atmospheric deposition (e.g., 1860) [Dentener, 2006], we estimate that about 75% and 88% of N deposition in 1993 and 2050, respectively, is attributable to human activity. Accordingly, the anthropogenic portion of emissions that we estimated using the IPCC methodology are ∼0.03–1.5 Tg N y−1 in 1993 and ∼0.6–3.0 in 2050. Contemporary anthropogenic N2O emission from rivers (0.68 Tg N y−1) [Beaulieu et al., 2011], are within the range of our estimates of deposition-related emissions from lakes. Again, these estimates are not strictly comparable because we do not consider all anthropogenic N inputs to lakes.
 In Colorado and elsewhere, N deposition has altered algal communities, water chemistry, and food web dynamics of lakes [Baron et al., 2000; Elser et al., 2010, 2009b]. Our data show that N deposition is associated with increased concentrations of NO3− and dissolved N2O, reduced CH4 production by sediments, and reduced concentrations of dissolved CH4. The microbial source of N2O is not clear, but enhanced nitrification and denitrification are likely. Regardless of background N loading rates, CO2 was the most dominant lake greenhouse gas. Enhanced N2O production, however, is particular concern because of its role in the destruction of stratospheric ozone. To our knowledge, this analysis constitutes the first global assessment of N2O emissions from lakes. Our high-level estimates suggest lakes should be considered in inventories of N2O emissions from aquatic ecosystems. Given that our approach is based only on atmospheric N loading, future efforts should consider N2O emissions related to all N inputs to lakes, including agricultural runoff and sewage. Indeed, emissions based on N deposition alone are 13–95% of total N2O emissions from rivers and estuaries. The inclusion of all N sources could substantially increase estimated N2O lake emissions. Additional studies are needed to constrain ratios of N2O-N:NO3− in lakes, considering the wide range we found among relatively few published studies (Table 6). More generally, further research in a variety of climates and across seasons is required to better understand lake N2O dynamics.
 We thank Erin Seybold, Marcia Kyle, and Melanie Engstrom for their assistance with field and laboratory work. David Huber provided valuable support for analyses of gas samples. We are also grateful to the staff of the Mountain Research Station, Niwot Ridge Long-term Ecological Research program, Rocky Mountain Biological Laboratory, Mountain Studies Institute, and Rocky Mountain National Park for their contributions in facilitating this study. The comments of two reviewers substantially improved the quality of this manuscript. Research support was provided by National Science Foundation grant DEB-0516494 to JJE and graduate student research grants to MLM from the Mountain Studies Institute, American Alpine Club, and Society of Wetland Scientists.