ECOTOXICOLOGY: A Comprehensive Treatment - Chapter 35 pdf

Clements: “3357_c035” — 2007/11/9 — 12:39 — page 771 — #1 35 Effects of Global Atmospheric Stressors on Ecosystem Processes 35.1 INTRODUCTION Global atmospheric stressors create widely distributed physical and chemical perturbations that impact the structure and function of ecosystems. In addition to their ubiquitous distribution and dif- fuse sources, global atmospheric stressors are especially problematic because they are not restricted by geopoliticalboundaries. In Chapter 26, we described the effects ofglobal atmosphericstressors on abundance, species diversity, community composition, and other structural characteristics. Here we will focus on descriptive and experimental studies that assess effects of increased CO 2 , N deposition, acidification, and ultraviolet radiation (UVR) on the function of aquatic and terrestrial ecosystems. 35.2 NITROGEN DEPOSITION AND ACIDIFICATION In Chapter 30, we described biogeochemical cycles and the important processes that control move- ment of elements in aquatic and terrestrial ecosystems. Significant increases in the global reservoirs of C, N, and S as a result of combustion of fossil fuels and agricultural/land use changes disrupt these natural cycles and have contributed to a variety of local and global environmental concerns. Global emissions of biologically reactive N compounds (e.g., NH 3 ,NH 4 , HNO 3 , and NO 3 ) have increased from about 15 teragrams (Tg) in 1860 to more than 165 Tg in 2000 (Galloway et al. 2003). Although effects of increased N deposition have not attracted the same attention from scientists and the public as other global atmospheric stressors such as chlorofluorocarbons (CFCs) and CO 2 , N poses serious threats to ecosystem processes. Potential negative effects of excess N on forest ecosystems were first described by Nihlgard (1985). Biologically reactive N compounds that accumulate in the atmo- sphere are rapidly deposited on the earth’s surface where they can affect net primary productivity (NPP) (Aber et al. 1995), disrupt N dynamics in soils (Gundersen et al. 1998), and contribute to eutrophication (Rabalais et al. 2002), acidification (Vitousek 1994), and subsequent loss of biolo- gical diversity (Stevens et al. 2004). In addition, N 2 O is a potent greenhouse gas that contributes to global climate change. Because the rates of production of reactive N in the biosphere greatly exceed rates of removal by denitrification, biologically active N rapidly accumulates in the environment. Predicting effects of N deposition on aquatic and terrestrial ecosystems is complicated by variation in regional climate, hydrologic characteristics, vegetation type, and other sources of anthropogenic disturbance (Aber et al. 2003). Assessing effects of N on ecosystems is also complicated because deposition often co-occurs with other stressors, such as heavy metals (Gawel et al. 1996). Finally, input and output of N are not necessarily coupled in all ecosystems, and leaching of nitrate will depend on nutrient status (Gundersen et al. 1998). 35.2.1 THE NITROGEN CASCADE Accumulation, transfer, and denitrification of biologically reactive N compounds through the biosphere and the changes that result have been termed the nitrogen cascade (Table 35.1) (Galloway et al. 2003). Forests and grassland ecosystems, especially the soil components, are major 771 © 2008 by Taylor & Francis Group, LLC Clements: “3357_c035” — 2007/11/9 — 12:39 — page 772 — #2 772 Ecotoxicology: A Comprehensive Treatment TABLE 35.1 Factors That Influence the Nitrogen Cascade in Aquatic and Terrestrial Ecosystems Ecosystem Accumulation Potential Transfer Potential Denitrification Potential Biological Effects Grasslands and forests High Moderate (high in some places) Low Biodiversity; NPP; mortality; groundwater Freshwater Low; higher in sediments Very high Moderate to high Biodiversity; altered community structure; eutrophication Coastal marine Low to moderate; higher in sediments Moderate High Biodiversity; altered community structure; algal blooms Source: Modified from Table 1 in Galloway et al. (2003). reservoirs for N. Because output of N in undisturbed forest and grassland ecosystems is generally quite low, residence time can be many years. Forest ecosystems are often N limited, and therefore N is cycled internally with little export to surface water, groundwater, or the atmosphere. As excess N deposition increases, other environmental factors will limit NPP and unused N leaches below the rooting zone, a process known as nitrogen saturation (Aber et al. 1995). Land use changes in forests and grasslands also have the potential to significantly alter internal N cycling and increase the export of N to aquatic ecosystems and the atmosphere. Fertilizers and runoff associated with agricultural and urban areas are the primary contributors of N to aquatic ecosystems and have been considered in previous chapters. However, nitrogen oxides (NO x ) from fossil fuel combustion account for about 25% of the reactive N in the environment. Although there is relatively little storage of N in overlying water, sediments represent a significant reservoir of N in aquatic ecosystems. Enrichment of aquatic ecosystems by N deposition can result in eutrophication, anoxia, and reduced biodiversity. Although natural aquatic ecosystems are highly retentive of N, this capacity for internal processing can be exceeded, especially in disturbed habitats, and downstream transport of N can contribute to eutrophication of coastal areas. Transport of N from rivers to coastal areas is generally regarded as one of the most serious threats to marine ecosystems (Rabalais et al. 2002). Accumulation potential of N in estuaries is relatively low; however, as with freshwater ecosystems coastal marine sediments may represent a significant N reservoir. Because of large amounts of organic material and low concentrations of dissolved oxygen in sediments, coastal marine ecosystems also have the greatest potential for conversion of reactive N to N 2 by denitrification, a process that is often enhanced by excess reactive N. In fact, denitrification in rivers and estuaries greatly reduces the amount of N transported from terrestrial to coastal and offshore areas (Galloway et al. 2003). 35.2.2 EFFECTS OF NDEPOSITION AND ACIDIFICATION IN AQUATIC ECOSYSTEMS Although most of the earlier studies of atmospheric deposition focused on ecosystem effects of sul- fur (S), attention in North America and Europe has shifted to concerns about effects of atmospheric N deposition. Atmospheric deposition of S and N from the mid-1980s to the mid-1990s show contrast- ing temporal patterns in North America (Sirois et al. 2001). While declines of SO 4 in precipitation were observed, concentrations of NO 3 and NH 4 generally remained constant or increased. These changes corresponded to decreased emissions of SO 4 and increased emissions of NO x over this same period. As previously N-limited forests became saturated, NO 3 is released to watersheds causing © 2008 by Taylor & Francis Group, LLC Clements: “3357_c035” — 2007/11/9 — 12:39 — page 773 — #3 Effects of Global Atmospheric Stressors on Ecosystem Processes 773 eutrophication, acidification and reductions in acid-neutralizing capacity (ANC; defined as the capa- city of the watershed to withstand strong acid inputs based on the difference between total cations and total anions). In fact, one of the most consistent indicators of N saturation in ecosystems is an increase in concentrations of NO 3 in stream water. Detecting these trends in streams requires access to long-term data that are often unavailable for many watersheds. Much of the research that describes biogeochemical responses of streams to N deposition has been conducted in Europe and the northeastern United States. In particular, experimental studies and long-term monitoring of SO 4 and NO 3 in stream water at Hubbard Brook Experimental Forest documented acidifica- tion effects on watersheds and significant decreases in base cation concentrations (Likens et al. 1996). Assuming that ecosystems can only tolerate a certain level of acidification before important processes become disrupted, defining the threshold point at which the capacity of an ecosystem to withstand additional inputs is exceeded is an important exercise. The concept of critical soil acidification load for a watershed is analogous to assimilative capacity in ecotoxicology. The idea assumes that once the net neutralizing capacity of an ecosystem is reached, additional atmospheric deposition will result in soil acidification. Critical soil acidification load is determined primarily by a balance between the weathering of base cations and leaching associated with deposition of SO 4 and NO x . Moayeri et al. (2001) developed a model to calculate critical soil acidification loads for a watershed in Ontario, Canada. Model results showed that soil acidification would occur faster in a harvested watershed compared to an old-growth watershed. Even relatively remote areas located away from sources of N can experience effects of N depos- ition. In general, water quality in alpine and subalpine lakes of the Rocky Mountains is relatively pristine, with low background concentrations of NO 3 (Williams and Tonnessen 2000). A survey of 44 high-elevation lakes (>3000 m a.s.l.) located on both sides of the continental divide in Color- ado indicated that higher NO 3 concentration and lower ANC of eastern lakes corresponded with greater atmospheric N deposition (Baron et al. 2000). Long-term changes in community composi- tion and productivity of diatoms, as revealed by paleolimnological records of lake sediments, were consistent with increased eutrophication. These increases in N deposition and shifts in diatom flora corresponded with increases in urban, agricultural, and industrial development on the Front Range of Colorado. Williams and Tonnessen (2000) used long-term monitoring data and synoptic surveys of 91 high-elevation lakes in the central Rocky Mountains to establish critical loads for inorganic N deposition. Episodic acidification of sensitive headwater catchments in remote Wilderness Areas has resulted from increased wet deposition of N. The concept of N saturation, originally developed to describe export of N in forest ecosystems, may also apply to watersheds. Similar to the acidification effects associated with S emissions, sensitivity of watersheds to N deposition will be influenced by underlying geology, hydrologic characteristics, soil type, and vegetation. Alpine and high-elevation watersheds, especially those located above treeline, may be especially susceptible to N deposition because of their relatively nonreactive bedrock, short growing season, and limited vegetation (Fenn et al. 1998). Williams et al. (1996a) reported that N deposition in the catchments of the Colorado Front Range was similar to that in other well-studied northeastern locations, including Hubbard Brook (New Hampshire) andAcadia National Park (Maine). A shift in nutrient dynamics from N-limited to N-saturated conditions was also observed in these high-elevation watersheds as a result of increased anthropogenic N deposition. Williams et al. (1996a) suggested that N saturation in high-elevation catchments may serve as an early warning of disruption in N cycling. Ecosystem responses of oligotrophiclakes to Ndeposition will also dependon nutrient conditions and the history of NO 3 availability. Nydick et al. (2004) measured effects of NO 3 enrichment in two alpine lakes with very different background levels of N. Enrichment significantly increased photosynthetic rate and chlorophyll a in a low N lake, but had no effects on a high N lake. Both NO 3 and PO 4 additions were necessary to increase productivity in the high N lake. Nydick et al. (2004) also reported that despite relatively little effect on benthic algal biomass, epilithon, surface sediment, © 2008 by Taylor & Francis Group, LLC Clements: “3357_c035” — 2007/11/9 — 12:39 — page 774 — #4 774 Ecotoxicology: A Comprehensive Treatment and subsurface sediment accounted for 57–92% of the NO 3 uptake, indicating the importance of benthic processes in these lakes. Ecosystem-level effects of SO 2 and NO x deposition on acidification of aquatic ecosystems has been studied extensively in Europe and North America (Hornung and Reynolds 1995, Likens et al. 1996, Schindler 1988). Much of the research conducted in the Experimental Lakes Area (ELA) focused on ecosystem processes, especially primary productivity (Schindler 1987, Schindler et al. 1985). Results of these and other studies of acidification showed reduced rates of primary and secondary production, decomposition, and nutrient cycling. Although the whole lake experimental studies conducted at Little Rock Lake in northern Wisconsin focused on community responses to acidification (Gonzalez and Frost 1994), Frost et al. (1999) speculated that loss of sensitive species and shifts in community composition would diminish the ability of acidified lakes to maintain system function. Effects of acidification on organic matter processing have been examined experimentally in natural and artificial streams. Burton et al. (1985) measured effects of acidification on decom- position of white birch and sugar maple in experimental stream channels (Figure 35.1). Reduced decomposition rates in acidified stream channels were attributed to lower density of macroinver- tebrates, particularly shredder caddisflies and detritivorous isopods. It is interesting to note that significant effects of acidification were not observed until relatively late in the study (>80 days), demonstrating the importance of long-term experiments. However, results of long-term experi- ments do not necessarily demonstrate significant ecological effects. Smock and Gazzera (1996) Percent dry weight remaining 30 40 50 60 70 80 90 Sugar maple Time (day) 0 50 100 150 200 250 300 30 40 50 60 70 80 90 100 White birch FIGURE 35.1 Decomposition rate (as percent dry weight remaining) of sugar maple and white birch in reference (closed symbols) and acidified (open symbols) stream channels. (Data from Table 2 in Burton et al. (1985).) © 2008 by Taylor & Francis Group, LLC Clements: “3357_c035” — 2007/11/9 — 12:39 — page 775 — #5 Effects of Global Atmospheric Stressors on Ecosystem Processes 775 introduced H 2 SO 4 to a low gradient, blackwater stream in Virginia (USA). Monthly additions over a 1-year period reduced benthic microbial respiration, but had no effect on processing rates of red maple leaves or abundance of macroinvertebrates. These results suggest that ecological effects of acidification on ecosystem processes will likely vary with location. Blackwater streams with nat- urally high levels of tannins and organic acids that typify this region are relatively insensitive to acidification. 35.2.3 EFFECTS OF NDEPOSITION AND ACIDIFICATION IN TERRESTRIAL ECOSYSTEMS Although negative effects of NO x deposition on lakes and streams have been frequently observed, changes in terrestrial ecosystems, especially forests, have received the most attention. Deposition rates of N toforest ecosystems range from2 kg N/ha/year inremote reference areas to 40kg N/ha/year in forests downwind of industrial sources (Aber et al. 1989). A large-scale survey of 68 grassland ecosystems across Great Britain showed a strong negative relationship between species richness and N deposition (Stevens et al. 2004). Although this paper focused on changes at the community level, it served to illustrate the widespread nature of this problem. At the current rate of N deposition in central Europe (17 kg N/ha/year), these researchers estimated that species richness was reduced by approximately 23% compared to grassland ecosystems receiving the lowest levels. Similar spatial gradients in N deposition and ecological effects are also evident in the northeastern United States. For example, relatively high rates of deposition have been measured in southern New York and Pennsylvania (12 kg N/ha/year), whereas low rates of deposition are measured in eastern Maine (<4 kg N/ha/year) (Aber et al. 2003). Although it is well documented that alpine and other high- elevation ecosystems are at greater risk from N pollution because of higher rates of deposition and increased sensitivity, other landscape factors that influence N deposition are not well understood. Direct measurement of wet and dry deposition rates in forest ecosystems is difficult. Weathers et al. (2000) developed a model to predict the influence of several landscape factors on N deposition in montane ecosystems. Usingconcentrations of Pb in forest floor soils as an index of N deposition, these investigators quantified effects of forestedges, elevation, aspect, and vegetation type on N deposition in montane forests. Fenn et al. (1998) published a comprehensive review of factors that predispose ecosystems to N saturation and the general ecosystem responses to N deposition. Because of the intimate linkages between forests and surrounding watersheds, research describing transport of N in terrestrial ecosys- tems has important implications for N transport to lakes and streams. For example, export of base cations, increased acidification, and elevated levels of nitrate and aluminum in streams are likely to be associated with N deposition in forests. Aber et al. (1989) provided a formal definition of the term N saturation and developed a hypothesized time course describing responses of forest ecosystems to N deposition (Figure 35.2). At a critical stage in this sequence of events, forests will likely become net sources of N rather than sinks. Aber et al. (1989) also described the limited capacity of some forest ecosystems to assimilate excess N and the potential interactions of N deposition with other atmospheric stressors such as ozone, sulfate, and heavy metals. 35.2.3.1 The NITREX Project One of the most comprehensive and spatially extensive experimental assessments of N deposition was conducted in several coniferous forests in northeastern Europe. The NITREX (nitrogen saturation experiments) project involved experimental additions of N to sites along a gradient of N pollution to examine changes in structural and functional characteristics (Emmett et al. 1998). Input rates of N ranged from 13 to 59 kg/ha/year across sites. Similar to the lotic intersite nitrogen experiment (LINX) experiments described in Chapter 30, an important objective of the NITREX project was to compare responses to N addition among sites. Experimental treatments involved both enhanced © 2008 by Taylor & Francis Group, LLC Clements: “3357_c035” — 2007/11/9 — 12:39 — page 776 — #6 776 Ecotoxicology: A Comprehensive Treatment Deposition begins Saturation Decline NPP Foliar biomass Foliar [N] Nitrate assimilation Relative units FIGURE 35.2 Predicted time course for changes in NPP, biomass, foliar N concentration, and nitrate (NO 3 ) assimilation in a forest ecosystem in response to chronic deposition of N. (Modified from Figure 1 in Aber et al. (1989).) N inputat sites with naturally low atmosphericdeposition and reduced N input (using exclusion roofs and an ion exchange system) at sites with high deposition. Researchers predicted that N addition in N-limited forests would have relatively little effect on leaching because of the high capacity for N retention in these systems. Most of the observed responses to N manipulations were consistent with expectations, with reductions in N status and NO 3 leaching occurring at sites where N deposition was reduced and increases occurring at sites where N deposition was enhanced (Gundersen et al. 1998). Shifts in N status and cycling rates following treatments generally supported the N saturation hypothesis (Aber et al. 1989). The most consistent responses to enhanced N deposition were changes in water quality, par- ticularly increased NO 3 . A strong relationship between N leaching and N status suggested that the distinction between N-limited and N-saturated forests could be quantitatively demonstrated (Gun- dersen et al. 1998). Nitrate leaching was observed within the first year of the experiments, whereas biological responses were often delayed. The strong link between N deposition and acidification was also demonstrated in the NITREX project. N additions caused a decrease in ANC, whereas experimental reductions in N caused an increase in ANC (Emmett et al. 1998). Ratios of carbon to nitrogen (C:N) were also good predictors of the onset of NO 3 leaching (Figure 35.3) because nitrification rates are stimulated as C:N declines, resulting in a decrease in the retention efficiency of N (Emmett et al. 1998). These results suggest a simple threshold response of N export. At C:N greater than 24, only a small proportion of nitrate leaches (approximately 10%); however, as C:N decreases, a rapid increase in the proportion of NO 3 leached was observed. Comparison of N dynamics at N-limited and N-saturated sites showed that microbial cycling of C and N was characterized by low NH 4 transformation and respiration rates at N-limited sites. Surprisingly, despite a wide range of variation in N deposition rates, N transformation showed relat- ively little variation. Gross mineralization and immobilization rates of NH 4 were highly correlated with respiration rates across ecosystems (Figure 35.4), indicating the important linkage between C and N dynamics and showing that simple measurements of CO 2 in soils could potentially serve as a measure of N cycling (Tietema 1998). The potential damaging effects of atmospheric deposition on forest productivity were clearly illus- trated by the NITREX project. Experimental reduction of N and sulfur inputs to a highly N-saturated site resulted in a 50% increase in tree growth (Emmett et al. 1998). However, in general, ecosystem responses to experimental manipulation of N were relatively modest and required longer periods © 2008 by Taylor & Francis Group, LLC Clements: “3357_c035” — 2007/11/9 — 12:39 — page 777 — #7 Effects of Global Atmospheric Stressors on Ecosystem Processes 777 C:N ratio of forest floor 15 20 25 30 35 NO 3out /NO 3in 0 1 2 3 FIGURE 35.3 Relationship between N leaching and carbon:nitrogen (C:N) ratios in the forest floor. Data are from experimental results of the NITREX project in northwestern Europe. (Modified from Figure 5 in Emmett et al. (1998).) Respiration (mg C/kg/day) 100 150 200 250 300 350 400 450 Gross N transformation (mg N/kg/day) 5 10 15 20 25 30 FIGURE 35.4 Relationship between gross N transformation rates and respiration in forests soils from the NITREX experiment. Solid circles = gross NH 4 mineralization rate; open circles = gross NH 4 immobilization rate. (Data from Tables 2 and 3 in Tietema (1998).) of time. Boxman et al. (1998) commented that it was “remarkable that no ecosystem components have responded with increasing vitality to the high N levels in the initially N limited, oligotrophic forests.” Mass loss in litter bags also increased along the gradient of N status, but effects of N manipulation were not significant. Responses of soil fauna to N treatments were generally less than initial differences among sites along the N gradient. The modest biological responses to N treatments may have resulted from the relatively short duration of these experiments. Gundersen et al. (1998) provided estimates of the amount of time required for several chemical and biological responses to N treatment (Table 35.2). In general, decomposition rates and changes in NPP were relatively slow processes compared with NO 3 leaching. These results again highlight the importance of conducting long-term manipulations for assessing ecosystem effects of N deposition. 35.2.3.2 Variation in Responses to N Deposition among Ecosystems Responses of forestecosystems toN deposition areoften variable, and factors thatcontrol differences in N retention and export are not completely understood. As described above, C:N ratios in soil are © 2008 by Taylor & Francis Group, LLC Clements: “3357_c035” — 2007/11/9 — 12:39 — page 778 — #8 778 Ecotoxicology: A Comprehensive Treatment TABLE 35.2 Predicted Timing of Selected Forest Ecosystem Responses to Changes in Chronic Additions and Reductions in N Deposition Pool or Process Response NO 3 leaching Fast Net mineralization Intermediate Decomposition Slow intermediate Denitrification Slow intermediate NPP Slow C:N of forest floor Slow C:N of mineral soil Very slow Fast = 1 year; intermediate = 2–4 years; slow ≥ 5 years. Note: Results are based on experimental manipulations of nitro- gen from the NITREX project. Source: From Table 8 in Gundersen et al. (1998). linked to thecapacity of forestecosystems to retainN, and decreasedC:N may occurunder conditions of chronic N deposition (Emmett et al. 1998). Surveys of wet deposition in old-growth stands of Engelmann Spruce showed elevated levels of NO 3 and NH 4 on the eastern side of the continental divide in Colorado (Baron et al. 2000). Greater N deposition was attributed to agricultural and atmospheric sources and was reflected in higher rates of mineralization and nitrification. Subsequent N fertilization experiments conducted in Engelmann Spruce forests on both sides of the continental divide showed that differences in soil conditions influenced responses to N treatments (Rueth et al. 2003). Mineralization rates were unaffected by N treatment on the western side of the continental divide but increased by approximately two times on the eastern side. Experimental studies ofN deposition at theELAin Ontario (Canada) havebeen conducted to con- trast responses ofdifferent ecosystem components.A2-yearN addition experiment (40 kgN/ha/year) was conducted in a boreal forest to compare effects in N-limited “forest islands” with naturally N- saturated lichen outcrops (Lamontagneand Schiff 1999). Responses to N treatments differed between habitat types. In contrast to forests, N-saturated lichen outcrops were highly sensitive to N depos- ition. After 2 years of treatment, lichen outcrops no longer retained additional N inputs, whereas the proportion of N retained in treated and reference forest–islands was similar (Figure 35.5). These data highlight the role that relatively small habitat patches in a landscape play in controlling ecosystem processes. Differences in abundance of dominant species will also complicate our ability to predict ecosys- tem responses to N deposition in forests. For example, species-specific differences in litter quality and rates of decomposition influence N cycling and availability. Surveys conducted along a gradi- ent of atmospheric N deposition in the northeastern United States showed that responses differed between tree species (Lovett and Rueth 1999). Rates of mineralization and nitrification of soils were significantly related to N deposition in maple plots but not in beech plots. Thus, while increased rates of mineralization and nitrification are typical responses to N deposition in forest ecosystems, the species composition of these forests should be considered when developing predictive models. Similarly, changes in community composition of forest ecosystems as a result of natural or anthropo- genic disturbance will affect responses to N deposition. Lovett and Rueth (1999) also demonstrated that although descriptive studies do not allow researchers to directly infer causality, comparison of ecosystem responses to chronic N deposition along a gradient can be a useful alternative to experimental treatments. © 2008 by Taylor & Francis Group, LLC Clements: “3357_c035” — 2007/11/9 — 12:39 — page 779 — #9 Effects of Global Atmospheric Stressors on Ecosystem Processes 779 Community type Bedrock Forest Retention coefficient 0.0 0.5 1.0 Reference watershed Treated watershed FIGURE 35.5 Responses of lichen bedrock communities and forest communities after 2 years of N additions in the ELA, Ontario, Canada. The figure shows retention coefficients ([Total N Input – Total N Output]/Total N Input). (Data are from Table 4 in Lamontagne and Schiff (1999).) 35.2.4 ECOSYSTEM RECOVERY FROM NDEPOSITION Mitigating the effects of N deposition on forest ecosystems will be challenging because of the widespread distribution of sources and tremendousvariation in ecologicaleffects. However, potential for recovery is relatively high if N deposition is significantly reduced. Experiments conducted in The Netherlands and Germany showed rapid improvements in water quality following reductions in N deposition to a watershed (Emmett et al. 1998). Increased tree growth was also observed at two of the three sites where N deposition was reduced (Boxman et al. 1998). Soils are a major sink for excess N in forests, and recent evidence suggests that microbial assimilation of NO 3 may be an important regulator of N retention. Because major reductions in N emissions are unlikely for some areas, management options should also include strategies to enhance the incorporation of N into soils (Fenn et al. 1998). Recovery of watersheds from the long-term effects of acidification may require many years if base cations are significantly depleted. Reductions in atmospheric deposition of sulfate as a result of the Clean Air Act have been associated with significant improvements in stream water chemistry. However, export of base cations, especially Ca and Mg, from acidified watersheds will likely delay recovery (Likens et al. 1996). Although the loss of base cations has been attributed primarily to prolonged exposure to acid rain and a decline in Ca in precipitation, ecological factors also influence Ca export. Hamburg et al. (2003) measured levels of Ca in the forest floor, abundance of snails (organisms that require Ca for growth), and Ca export in stream water from hardwood forests of various ages. Results showed that Ca concentration in snails, litterfall, and the forest floor and export of Ca in stream water increased with forest age. Calcium mobilization in young stands (4.6–6.0 g Ca/m/year) was much greater than in old stands (0.4 g Ca/m/year), indicating that forest aging significantly influenced Ca dynamics. Recent amendments to the 1990 Clean Air Act in the United States are expected to have signi- ficant effects on air quality and water chemistry across large broad geographic regions. Assessing these changes will require integrated studies of physicochemical and biological responses over large regional areas andfor relatively long periodsof time. Because long-term monitoring dataat a regional level are generally lacking, researchers are often required to use trends from site-specific results to infer regionalpatterns. Stoddard et al. (1998) analyzed trends in waterchemistry from 44 Adirondack and New England (USA) lakes that were sampled from 1982 to 1994. Long-term trends in meas- ures of acidic deposition (SO 4 and NO 3 concentrations in stream water) and watershed responses to acidification (ANC and export of base cations) differed between subregions. In particular, ANC increased overtime in New England lakes but decreased inAdirondack Lakes. These results were not © 2008 by Taylor & Francis Group, LLC Clements: “3357_c035” — 2007/11/9 — 12:39 — page 780 — #10 780 Ecotoxicology: A Comprehensive Treatment expected and indicate that potential for recovery from acid deposition would be considerably less in the Adirondacks. The most significant finding of this research is that even with long-term data and a solid mechanistic understanding of physicochemical relationships, predicting regional trends based on well-defined subpopulations of sentinel lakes is difficult (Stoddard et al. 1998). The challenges of making accurate regional predictions on the basis of a very well-understood phenomenon highlight the potential difficulties associated with quantifying effects of poorly defined stressors such as CO 2 and UV-B radiation. 35.3 ULTRAVIOLET RADIATION to really demonstrate UV-B radiation impacts at the ecosystem level requires establishing a chain of cause and effect from molecule to ecosystem. (Bassman 2004) Ecosystem-level experiments are the only method of detecting UV influences on the myriad of competitive and trophic interactions present in nature. (Flint et al. 2003) 35.3.1 AQUATIC ECOSYSTEMS Effects of UVR on ecosystem processes have been studied extensively, especially in pelagic mar- ine and lentic systems. As a result of this comprehensive research effort, there exists sufficient information concerning effects of UVR to develop reasonably detailed ecological risk assessments for certain groups of organisms and processes (Hansen et al. 2003). Several excellent reviews on the effects of UVR have been published recently. Day and Neale (2002) have provided the most comprehensive treatment of UV-B effects in aquatic and terrestrial ecosystems, with an emphasis on primary producers. One of the most consistent observations in studies of UV-B effects on marine and freshwater phytoplankton is reduced primary production (Kinzie et al. 1998, Mostajir et al. 1999, Neale et al. 1998, Smith et al. 1992, Williamson 1995). UV-B causes damage to Photosystem II, and effects of UVR on primary production and other ecosystem processes can be extensive. Gala and Giesy (1991) estimated that exposure to UV-B reduced primary production of a natural assemblage of phytoplankton from Lake Michigan by 25%. Exposure to simulated UV-B reduced photosynthesis by 40% in Georgian Bay (Furgal and Smith 1997). Remarkably, even short-term exposure to surface radiation (e.g., 30 min) can be sufficient to inhibit photosynthesis (Marwood et al. 2000). Because of proximity to the ozone depletion zone and the intense exposure to UVR during the early australspring (October–November), considerableresearch effort has focused on phytoplankton in Antarctic waters, where a 50% reduction in ozone has been documented. Smith et al. (1992) conducted transects in the Antarctic marginal ice zone and reported a 6–12% reduction in primary production associated with ozone depletion during a 6-week cruise. Assuming that this reduction is representative of the entire area and integrating results over the austral spring, Smith et al. (1992) concluded that this change corresponded to an approximately 2% reduction in annual production of the Southern Ocean. Similar experiments conducted in the Weddell Sea showed that productivity of marine phytoplankton decreased as a cumulative function of UV exposure, indicating little evidence of photorepair (Neale et al. 1998). These researchers also documented considerable variation in sensitivity of primary production among sites, and attributed this variation to UV exposure before sampling. 35.3.1.1 Methodological Considerations With respect to experimental methodology, most aquatic investigations involved the removal of different wavelengths of UVR using various filters, although some have employed experimental © 2008 by Taylor & Francis Group, LLC [...]... radiation that simulated a 14% reduction in ozone depletion Photosynthesis and respiration were either similar or increased slightly under moderate UV-B exposure These researchers speculated that alpine plants are acclimated to UV-B and that photosynthetic processes are protected by morphological adaptations such as increased leaf thickness Meta-analysis offers a quantitative approach for integrating results... on many aquatic ecosystems Wetland loss in boreal regions of Alaska and Canada is likely to result in additional releases of CO2 into the atmosphere Coastal wetlands are particularly vulnerable to sea level rise associated with increasing global temperatures Most specific ecosystem responses to climate change cannot be predicted because new combinations of native and nonnative species will interact... stream discharge are available for many watersheds (e.g., U.S Geological Survey); therefore, predictive models that relate regional changes in climate to altered flow regimes within a watershed can be developed Associating local weather patterns and stream discharge over the past several decades may also provide useful insights into potential trends associated with global climate change However, extrapolation... to assess complex relationships among variables This approach is especially appropriate for assessing terrestrial ecosystem responses to UV-B because effects are expected to be relatively subtle and often indirect Searles et al (2001) conducted meta-analysis of 62 papers that investigated effects of UV-B radiation on the concentration of UV-B-absorbing compounds, growth, morphological variables, and... factors, and quantification of effects on ecosystem processes is often challenging (Marwood et al 2000) In addition to the elevated UV-B levels in the Southern Ocean, increases in UV-B radiation occur at higher elevations (increasing by approximately 20% for each 1000 m), placing alpine and subalpine ecosystems at considerable risk (Blumthaler and Ambach 1990, Sommaruga 2001) Alpine lakes and streams... experimental techniques used to enhance or reduce UVR were described in Chapter 26 In particular, there has been considerable discussion of the artificial wavelength spectra produced by UV lamps Unrealistic combinations of UV -A, UV-B, and photosynthetically active radiation (PAR) may artificially enhance effects of UV-B (Caldwell and Flint 1997) These problems may be partially addressed by either calculating... STRESSORS AND CONTAMINANTS Interactions between global atmospheric stressors and other contaminants have been reported in several aquatic ecosystems In general, reduced or altered streamflow patterns as a result of lower precipitation may concentrate chemicals in streams receiving pollutants (Carpenter et al 199 2a) Some contaminants may become more bioavailable under conditions of increased temperature or acidification... United States (100–150 kg/ha), but there was considerable annual variation (0.2 Pg C uptake) Because of this high annual variability, any 2- to 3-year period would be insufficient to assess potential C storage, thus highlighting the need for long-term assessments (Schimel et al 2000) These researchers also note that effects of forest management and agricultural abandonment are similar... past 100 years in several ecosystems (Graybill and Idso 1993, Nicolussi et al 1995) Braswell et al (1997) measured CO2 , temperature anomalies, and vegetation to examine the response of vegetation to temperature variability at a global scale Correlations among vegetation growth, temperature, and CO2 were lagged, suggesting that responses are regulated by biogeochemical feedbacks Patterns observed at... vegetation such as grasslands, alpine meadows, and peatlands (Price and Waser 2000, Shaw and Harte 2001, Shaw et al 2002, Updegraff et al 2001) Subalpine vegetation in the Rocky Mountains showed relatively little response to warming over a 4-year period (Price and Waser 2000) These researchers speculated that despite earlier snowmelt in treated plots, drying soils limited photosynthesis and microbial activity . Hampshire) andAcadia National Park (Maine). A shift in nutrient dynamics from N-limited to N-saturated conditions was also observed in these high-elevation watersheds as a result of increased anthropogenic. Global Climate Change • Aquatic and wetland ecosystems are very vulnerable to climate change. • Increases in water temperature will cause a shift in the thermal suitability of aquatic habitats. that relate regional changes in climate to altered flow regimes within a watershed can be developed. Associating local weather patterns and stream discharge over the past several decades may also