Historical increases and declines in nitrogen emissions recorded across the

Summary

Human activities have significantly increased the amount of biologically available nitrogen (N) since the industrial revolution with global atmospheric N deposition increasing three-fold since the mid-19^th century (Galloway et al., 2004; Rockström et al., 2009). Increased N deposition can cause acidification, eutrophication, forest decline, biodiversity loss, and degraded water quality (Davidson et al., 2012; Galloway et al., 2003). Understanding changes in N deposition is limited by spatially coarse monitoring networks that have documented N deposition since the late-20^th century (Fenn et al., 2009). Here we use herbaria lichen specimens as biological indicators that integrate N deposition and extend temporal and spatial records in the western United States (Fig.

1). Lichen N content documented increases in N deposition during the early- and mid-20^th century with the expansion of agriculture, industry, and fossil fuel use, and decreases at the turn of the 21^st century with the implementation of federal regulations controlling oxidized N emissions in the United States (US Environmental Protection Agency, 2000). Spatial variation in this pattern indicates regional differences in N deposition variation over time with the greatest increases in the Mediterranean California region where the most urban and significant agricultural development occurred. The North American Deserts region where lichen specimens were collected near major agricultural areas also documented greater increases in N deposition than regions with less urban and agricultural development. Previously established thresholds (Geiser, 1996) for lichen N content indicate when and where elevated N deposition may have negatively impacted ecosystems with more exceedances observed in the Mediterranean California region than in the Northwestern

Forested Mountains. Our results provide a way to assess historical impacts of elevated N deposition prior to the establishment of monitoring networks and demonstrate the effectiveness of regulations to decrease N deposition. The observed trends of decreasing N deposition however are not guaranteed to continue since current regulations focus on oxidized N emissions while emissions of reduced N compounds are still unregulated and increasing (Du et al., 2014).

Introduction

Excess N from industrial, fossil fuel, and agricultural emissions enters ecosystems as atmospheric N deposition, triggering a cascade of changes in terrestrial, freshwater, and marine ecosystems (Davidson et al., 2012; Galloway et al., 2004, 2003). Studies of elevated N deposition and its impacts have focused primarily on areas of high, chronic N deposition in the eastern United States (US) and Europe (Holland et al. 2005). The western US, historically, has been thought to have overall low N deposition, but localized areas of highly elevated N deposition rates occur, and some ecosystems in the western US are sensitive to even slight increases deposition above background levels (Fenn et al., 2003a, 2003b).

Despite, hot spots of elevated N deposition greater than 25 kg N ha^-1 yr^-1 and highly sensitive ecosystems, the distribution of N deposition monitoring networks is sparse across the large, heterogeneous landscapes of the western US; the National Atmospheric Deposition Program (NADP) and the Clean Air Status and Trends Network (CASTNET) have only 73 of 267 and 25 of 95 sites, respectively, across the eleven states in the western US (Lamb and Bowersox 2000, Fenn et al. 2009, Cummings et al. 2014). Additionally, they have only measured wet deposition since 1978 and dry deposition since 1990 (Lamb and Bowersox 2000, AMEC Environment &

Infrastructure 2015). This leaves undocumented the changes in N deposition levels associated with population growth, economic development, agricultural intensification, and environmental legislation that occurred earlier in the 20^th century (US Environmental Protection Agency 2000).

We analyzed herbarium and field specimens of two epiphytic lichen species Alectoria sarmentosa (Ach.) (hereafter “Alectoria”) and Letharia vulpina (L.) Hue (hereafter “Letharia”) for lichen N content (%N) and lichen N stable isotope composition (δ¹⁵N) to identify changes in N deposition and to fill in gaps in the sparsely monitored western US (Fig. 1). Lichens are ubiquitous and sensitive indicators of N deposition (Palmqvist et al., 2002; Pardo et al., 2011). Lichen %N is strongly correlated with throughfall N deposition and provides reliable estimates of N deposition rates under both high (Boltersdorf and Werner, 2014) and low (McMurray et al., 2013; Root et al., 2013) levels of deposition.

We coupled lichen %N measurements to measurements of lichen δ¹⁵N to better understand sources contributing to the changes observed in deposition amounts (Chapter 1). The δ¹⁵N of reduced (NHx) emissions is generally lower than the δ¹⁵N of oxidized (NOx) emissions from the same type of source, although the range of δ¹⁵N values often overlaps (Chapter 1 Fig. 2). Agricultural emission sources of NOx, including crop fields and concentrated animal feeding operations, vary between −19.8‰ and −48.6‰ and are lower than NOx from industrial sources which vary between +9.0‰ and +25.6‰ (Felix et al., 2013, 2012; Felix and Elliott, 2014; Li and Wang, 2008). Vehicle NOx emission span these ranges with δ¹⁵N ranging from −19.9‰ to +15.0‰ however in studies directly comparing δ¹⁵N from different emission sources, vehicle emissions are consistently higher than agricultural emission sources for both NOx and NHx (Felix et al., 2013; Felix and Elliott,

2014; Walters et al., 2015). The δ¹⁵N of NHx emissions from industrial sources ranges from

−11.3‰ to −14.6‰ and in direct comparisons of emission sources, is higher than agricultural emissions which range from −22.8‰ to −56.1‰ (Felix et al., 2013). Other studies, however, show higher δ¹⁵N from agricultural emission sources up to −1.3‰, although these measurements could have had mixing with other emission sources due to sampling location (Felix et al., 2014). The δ¹⁵N of deposition and lichens reflects the composition of emission sources with studies showing decreasing δ¹⁵N indicating greater contributions from agriculture and increasing δ¹⁵N indicating greater contributions from industrial and fossil fuel combustion emission sources (Bermejo-Orduna et al., 2014; Boltersdorf and Werner, 2014).

Our goal is to reconstruct N deposition amounts through lichen %N and emission sources through lichen δ¹⁵N in the western US during the past century by using lichen herbaria specimens as biological indicators. This will provide insight into historic N deposition levels, increases in deposition with the growth and development of cities and agriculture, and the duration of ecosystem impacts from elevated N deposition. We hypothesize that lichen %N increases over time as N emissions and deposition increase due to anthropogenic sources including industry, fossil fuel combustion, and agriculture. We also hypothesize that lichen δ¹⁵N has more complicated patterns over time dependent on the different types of emission sources located nearby.

Results and Discussion

Our measurements of specimens from the early 20^th to 21^st century provide the first continental scale indicators of how urban development and agricultural intensification (Fenn et al., 2003b) drive increases in atmospheric N deposition across the western US (Fig. 2, Fig. 3). Lichen %N did

increase over time and was significantly higher (p<0.05) in the 1980s for Letharia at 1.15% (Table 1) and in the 1990s for Alectoria at 0.54% (Table 2) than at other times. Lichen %N correlates extremely well with deposition levels across various species (Boltersdorf et al., 2014; McMurray et al., 2013; Root et al., 2013), and the lichen %N of both species reflects the increase in N emissions and subsequent deposition across the western US driven by population growth and the subsequent increases in fossil fuels use and agricultural production. Human populations grew 15-fold in the western US during the 20^th century from nearly 4.1 million in 1900 to greater than 61 million people in 2000 (Forstall, 1996; Mackun et al., 2010). Agricultural production also expanded and intensified with N fertilizer use, a proxy for agricultural production and intensification, increasing from 7.7 million kg N in 1945 to over 1.6 billion kg N in 1985 across the western US (Alexander and Smith, 1990). The lichen δ¹⁵N had more complicated trends with different patterns for Letharia and Alectoria, but for both species, it still can begin to identify sources of these changes in lichen %N. Multiple types of sources including both fossil fuel combustion and agricultural intensification likely contributed to the changes in Letharia %N as the δ¹⁵N of Letharia did not change in a consistent direction as Letharia %N increased (Fig 3, Table 3). Agricultural intensification likely contributed more to increases in Alectoria %N, however, due to decreasing Alectoria δ¹⁵N as %N increased (Fig 3, Table 4).

The increase in N deposition documented here through lichen %N corresponds to previous studies detecting increased N deposition beginning in the mid-20^th century. Lake sediment cores across the western US from Washington (Sheibley et al., 2014), Utah (Hundey et al., 2014), Colorado (Wolfe et al., 2001), and Wyoming (Spaulding et al., 2015) document significant increases in N deposition as recorded by shifts in the composition of diatom species. These shifts begin to occur

between the 1950s and the 1970s and align with when lichen %N began to increase in both Letharia and Alectoria. Ice cores from western US glaciers also document the same temporal pattern in N deposition as lichens and lake sediments with significant increases in N deposition being recorded just prior to 1970 (Naftz et al., 2011). Even outside the US, a similar pattern for significant increases in N deposition beginning during the mid-20^th century are documented in remote lake sediment cores (Holtgrieve et al., 2011) and Greenland ice cores (Geng et al., 2014;

Hastings et al., 2009).

Regional variation exists throughout the western US in the timing and degree of increase in N deposition and corresponding lichen %N. This corresponds well to the variation in major increases in N deposition documented by shifts in diatom species in lake sediment cores which began between 1950 (Hundey et al., 2016; Wolfe et al., 2001) and 1970 (Sheibley et al., 2014) depending on region. The greatest increases in lichen %N documented here occurred for Letharia %N in the Mediterranean California region which had significantly higher %N than other regions (Table 1).

Increases in Letharia %N in the Mediterranean California coincide with the greatest rates of urban development and agricultural intensification as the human population increased from ~1.5 million to nearly 34 million people during the 20^th century (Forstall, 1996; Mackun et al., 2010) and N fertilizer use increased 10-fold from 63.5 million kg N to 655 million kg N between 1945 and 1985 (Alexander and Smith, 1990) in California (Fig. 2). Differences in Letharia δ¹⁵N between time periods were not significant in the Mediterranean California region despite the dramatic changes in Letharia %N (Table 3). The lack of change in Letharia δ¹⁵N (Fig. 3, Table 3) indicates multiple types of emissions sources including urban, industrial fossil fuel emissions with higher δ¹⁵N and

agricultural emissions with lower δ¹⁵N are contributing to the significant increases in Letharia %N (Fig 2, Table 1).

Letharia in the North American Deserts region had similar patterns as in the California Mediterranean region with no significant change in Letharia δ¹⁵N and a significant but not as large increase in Letharia %N between the 1930s and the 1980s (Fig. 2, Table 1). Specimens from the North American Deserts region were primarily collected around the Columbia River Basin in Washington (Fig. 1), an area of prodigious agricultural production (Fenn et al., 2003b; Jennings et al., 1990). Washington had major increases in N fertilizer use from almost 6 million kg N yr^-1 to over 234 million kg N yr^-1 between 1945 and 1985 (Alexander and Smith, 1990). However, both industrial and fossil fuel emission sources likely contributed to this increase in Letharia %N as Letharia δ¹⁵N did not change significantly (Fig. 3, Table 3). An increase solely from agricultural emissions would be expected to decrease the δ¹⁵N of Letharia (Boltersdorf et al., 2014; Boltersdorf and Werner, 2014), but that is not what we observed. Most of the growth in industrial and fossil fuel emissions in the Pacific Northwest occurred along the Interstate 5 corridor on the west side of the Cascade Mountains which is outside of but upwind of the Columbia River Basin and the North American Deserts region (Cummings et al., 2014). Atmospheric transport could have brought emissions from these sources to the lichen samples collected in the Columbia River Basin of the North American Deserts region.

Trends for Letharia %N were similar for the Marine West Coast Forest and Northwestern Forested Mountains regions, however with a smaller amplitude than the California Mediterranean and North American Deserts regions. The Marine West Coast Forest and Northwestern Forested Mountains

regions are more removed from the major urban and agricultural areas which likely contributed to the smaller increases in Letharia N% over time compared to the other regions containing larger urban and agricultural areas. Letharia δ¹⁵N varied significantly in the Northwest Forested Mountains, but in both regions, Letharia δ¹⁵N, again, indicates both industrial, fossil fuel emission sources and agricultural emission sources drove changes in Letharia %N with no significant changes in Letharia δ¹⁵N in the Marine West Coast Forest region and no directional changes over time in Letharia δ¹⁵N in the Northwestern Forested Mountains region (Table 3).

Trends for Alectoria %N were similar to Letharia %N with a significant increase from the 1930s to the 1990s although the magnitude of change in Alectoria %N was smaller than Letharia %N (Fig. 2, Table 2). Regionally, the increase from the early to late 20^th century was not significant.

Unlike Letharia, the Alectoria δ¹⁵N decreased significantly (Table 4) as Alectoria %N increased, indicating agricultural emission sources as primary contributors to changes in deposition throughout Alectoria’s range of the Pacific Northwest where several areas of significant agricultural production exist (Cummings et al., 2014). A decrease in the δ¹⁵N of deposition has also been documented throughout the 20^th Century in remote locations of North American and Greenland through lake sediment cores (Holtgrieve et al., 2011) and ice cores (Geng et al., 2014;

Hastings et al., 2009). Alectoria δ¹⁵N illustrates this same decrease with Alectoria specimens collected from more remote regions with fewer major N emission sources. Letharia δ¹⁵N, however, does not decrease but remains relatively consistent over time, likely due to Letharia specimens originating in regions with greater numbers of major N emission sources. The more remote, pristine location of Alectoria and the other proxiesmay contribute to the observed patterns in δ¹⁵N

of deposition via proxies such as lichens and diatom community composition in lake sediment cores.

Differences between Alectoria and Letharia in %N and δ¹⁵N within regions may highlight local variation in N deposition. Alectoria occurs at sites with more moisture and is more sensitive to increases in N deposition than Letharia (Geiser, 1996). The orographic effects on precipitation driven by the Olympic and Cascade Mountain ranges create variation in moisture across the region.

Dry and moist locations could be experiencing different levels of N deposition that would account for observed differences between species. In addition, the greater sensitivity of Alectoria to increased N deposition may cause this species to be extirpated more quickly with a limited ability to document increases in N deposition. Differences between the two species in habitat and sensitivity to elevated N deposition helps to observe a broader array of ecosystems across the western US and different degrees of change in N deposition fluxes. Letharia adds the Mediterranean California and North American Deserts region. Co-occurrence of Letharia and Alectoria documents within region variation in addition to between region variation and by recording more significant increases in lichen %N in areas where Letharia occurred than in areas where Alectoria occurred. Both the Marine West Coast Forest and Northwestern Forested Mountain regions show significant increases in Letharia %N (Table 1) but not Alectoria %N (Table 2) and demonstrate this within region variability in N deposition between the habitats of these two species. Analyzing the composition of multiple indicator species helps counter limitations on the occurrence of a single species due to geographic range or pollutant tolerance and reveals a more extensive and more detailed picture of N deposition across the western US.

Critical loads (CL) are levels of pollution above which ecosystems experience negative impacts (Pardo et al., 2011), and their use here provides region-specific evidence for the ecological impacts in the western US. Lichens are the most sensitive terrestrial indicator of CLs (Pardo et al., 2011) and indicate the first terrestrial impacts caused by N deposition. Species-specific CL thresholds of lichen %N have been established by the US Forest Service for the Pacific Northwest (Fenn et al., 2011; Geiser, 1996; Pardo et al., 2011). Numeric CL thresholds vary between regions due to differences in climate, in particular precipitation, and its effect on lichen physiology (McMurray et al., 2015). The CL thresholds used here (Geiser, 1996) were developed specifically for Washington and Oregon lichens and are therefore not necessarily the exact CL value for other regions; they do, however, provide insight into the extent and duration of potential impacts from elevated N deposition. Mean western US Letharia %N exceeded the CL threshold from 1971 to 2000 with regional mean Letharia %N exceeding the CL threshold beginning in the 1930s in the California Mediterranean region and in the 1960s in the Marine West Coast Forests and North American Deserts regions. Given the lack of directional change in Letharia δ¹⁵N, these CL exceedances are likely driven by multiple types of emission sources including agricultural, industrial, and fossil fuel combustion. Chronically elevated N deposition can compound ecological impacts over time (Galloway et al., 2003). Initial impacts are often to the most sensitive species in terrestrial and aquatic habitats including lichens and diatoms, but over time, continuously elevated N deposition affects more tolerant species, decreases biodiversity, and increases N losses from the ecosystem (Fenn et al., 2011). Chronically elevated N deposition has caused altered plant, algae, and soil microbial communities in the Rocky Mountains of Colorado and caused N losses from forest ecosystems and decreased soil fertility throughout central and southern California (Davidson et al., 2012; Fenn et al., 2011, 2003a; Jovan and McCune, 2006;

Pardo et al., 2011). Decades-long CL exceedances should extirpate the sensitive lichen species used here. There are two reasons this may not have occurred. First, CL thresholds for these regions may be higher than those developed for Washington and Oregon. Second, lichens likely persisted where N deposition was near or above CL thresholds just prior to collection. Sampling locations varied within regions between time periods due to the nature of herbaria specimen collections which means sampling could have occurred in progressively remote areas as they were the only places with any lichens of these species still remaining. The long CL exceedances documented here indicate chronically elevated N deposition driven by multiple types of emission sources and suggest compounding and expanding ecological impacts across the western US throughout the 20^th Century.

More local scale impacts from elevated N deposition also are likely across all regions, even if mean lichen %N did not exceed CL thresholds. Mean Alectoria %N in the Marine West Coast Forest and Northwestern Forested Mountains regions along with mean Letharia %N in the Northwestern Forested Mountains region remained below the CL thresholds (Fig. 2) indicating lower N deposition throughout these regions. Some observations within these regions, however, exceeded established thresholds. Local exceedances highlight variability in N deposition fluxes across the landscape and the potential for hot spots of high N deposition. Variability in N deposition across the western US results in variable impacts from N deposition across the landscape (McMurray et al., 2015). Impacts to lichen community composition across the Pacific Northwest, for example, are highly variable across the region with sites dominated by nitrophilous lichen species and sites dominated by pollution-sensitive lichen species often interspersed among each other (Fenn et al., 2003a). Differences between Alectoria and Letharia within a region also

highlight local variation in N deposition. The difference in the Marine West Coast Forest—where Letharia exceeds the CL thresholds and Alectoria does not—could be driven by differences in habitat and sensitivity to N deposition in ways similar to %N patterns. The high variability in N deposition across regions documented here and in other studies (Fenn et al., 2003a; McMurray et al., 2015) deserves more recognition, as many monitoring and modelling efforts often use coarse spatial resolution that homogenizes the heterogeneous flux of N from deposition into ecosystems with implications for understanding impacts at finer spatial scales. Studies comparing the range of modelled and measured N deposition in the northern Rocky Mountains show a modelled range half that of the measured range of N deposition rates (McMurray et al., 2015).

Other factors including climate and other pollutants like sulfur deposition and ozone could also contribute to the differences in %N, δ¹⁵N, and CL exceedances between lichen species (Geiser and Neitlich, 2007; Jovan and McCune, 2005; McMurray et al., 2015). Letharia is more tolerant of drier climates and of sulfur pollution than Alectoria (Geiser, 1996), and effects of ozone on these species are unknown (Jovan and McCune, 2005). Teasing apart the driver of a species absence can be difficult since multiple culprits from N to other pollutants to climate or an interaction between such factors could stress the lichen species and limit their occurrence both within and between regions (Blett et al., 2003; Jovan and McCune, 2005). Stressed lichens may not be able to assimilate N deposition as well as healthy lichens which could hinder their reliability as N deposition indicators (Munzi et al., 2009). In addition, levels of different pollutants often follow similar spatial patterns making it difficult to isolate a single driver from field samples (Jovan and McCune, 2005).

Chronic and localized increases in N deposition have the potential to impact ecosystems but are rarely accounted for in N deposition studies (Lu et al., 2011) where ambient levels are used as a control, and a pristine status is assumed in the western US (Pardo et al., 2011). Historical exceedances observed here demonstrate that this assumption may not be valid and at a minimum should be questioned especially as N addition studies are reviewed and synthesized. Studies of

In document ATMOSPHERIC NITROGEN DEPOSITION IN THE WESTERN UNITED STATES: SOURCES, SINKS AND CHANGES OVER TIME SARAH MARIE ANDERSON (Page 48-81)