Kevin E. Percy
It is estimated that 49% of forests (17 million km2) will be exposed to damaging concentrations of tropospheric O3 by 2100. Global forest area at risk from S deposition exceeding critical loads may reach 5.9 million km2 by 2050. Coincidentally, the world climate is changing. Process-level changes in North American forests due to air pollutants have been documented for diverse forest types. In Europe, critical loads for acid deposition and critical levels for ozone are exceeded in most forested areas measured. Some 22% of trees rated in 2001 were moderately to severely defoliated. It is unclear at present whether climate change will lessen or enhance air pollution effects on forest ecosystem function and health.
According to El-Lakany, "Never before have the earths ecosystems been so greatly affected by our presence" (FAO 2001). In 2000, the worlds forests covered 3,869 million ha, or about 30% of the worlds land area (Figure 1). The net change in forest area between 1990-2000 was -9.4 M ha yr-1 with most of the losses in the tropics. The forests of Europe (mainly boreal forests in Scandinavia and the Russian Federation) and North/Central America represented 27% and 14% of world forests, respectively. These forests increased (Europe +0.9 M ha yr-1) or decreased (North/Central America -0.6 M ha yr-1) to a much lesser extent than tropical forests during that period (FAO 2001).
Figure 1. Distribution of world forests in 2000 (FAO, 2001)
The role of air pollutants play in forest health was recently reviewed (Percy 2002). At continental to national scales, health is at risk from tropospheric ozone (O3), and wet/dry acid deposition. Rising atmospheric carbon dioxide (CO2) concentrations may be beneficial in the short term, but new work demonstrates that "bottom-up" effects may negatively affect ecosystem function over the longer term (Karnosky et al. 2002). Less certain is the emerging risk posed by fine particulate matter (PM 2.5), persistent organic pollutants, and ultraviolet-B radiation.
Against this backdrop of changing atmospheric chemistry, the worlds physical climate is changing. Global average surface temperature has increased 0.6 ±0.20C since the late 1800s and annual land precipitation has increased in the middle/high latitudes in the Northern Hemisphere (Houghton et al. 2001). Relatively modest increases in global temperature are possible under some emissions scenarios (Stott and Kettleborough 2002).
Tropospheric O3 is the most pervasive air pollutant affecting forests. Levels were about 10-15 ppb a century ago, compared with 30-40 ppb measured as background around the world today (Finlayson-Pitts and Pitts 1999). Daily maximum 1-hour O3 concentrations in the US have decreased by 10% (EPA 1997). Data (Dann 2001) for Canada indicate O3 levels increasing, resulting in a 4th highest daily (3-year average) maximum concentration in 1999 of >70 ppb O3. Monitoring in Europe shows episodes of O3 occur over the continent every summer (Hjellbrekke and Solberg 2002). By 1996, 90% of pan-European monitoring plots exceeded critical levels. In 2000, exceedance (AOT40> 10,000 ppbh) was considerable over large parts of central and Eastern Europe. Ozone concentrations in many regions of Europe are sufficient to adversely affect tree growth (Matyssek and Innes 1999).
Fowler et al. (1999) report sulfur dioxide (SO2) emissions have declined in North America (-25%) and Europe (-48%) since the 1980s. For Europe, average reduction was 62.6%, compared with 40% in Canada, and 21% in the US (EMEP 2000). In 1996, critical levels for SO2 were exceeded at 20% of European Level II monitoring plots (Muller-Edzards et al. 1997). Data from air sampling in 2000 (Hjellbrekke 2002) indicated SO2 levels highest (> 2. µg m-3) in central and eastern Europe.
Nitrogen dioxide (NO2) emissions have not been reduced to same degree. For Europe, average reduction in emissions was 38%. Data for 2000 (Hjellbrekke 2002) show highest NO2 levels in Belgium and the Netherlands (>5.0 µg m-3), with elevated levels (2.0-5.0 µg m-3) in much of Europe. Anthropogenic NO2 emissions in the US decreased 2%, while NO2 emissions in Canada increased 4% between 1980-1998 (EMEP 2000).
Concentrations of atmospheric carbon dioxide (CO2) rose from 280 ppm in the pre-industrial era to 367 ppm today; doubling to 560 ppm is anticipated (Houghton et al. 2001). Forests can reduce the rise in CO2, as they are carbon limited at current CO2 levels (Drake et al. 1997). However, knowledge gaps remain (Karnosky et al. 2001a), including the potential of O3 to reduce productivity gains (Isebrands et al. 2001; Karnosky et al., 2002; Percy et al. 2002). For insight into CO2 and forests, the reader is referred to the treatise by Karnosky et al. (2001b)
In northeastern hardwood forests, sugar maple branch dieback was correlated with exceedance of S/N critical loads (Arp et al. 1996). Acid deposition has accelerated soil cation loss (Foster et al. 1992), and long-term depletion has led to predictions of calcium deficiency (McLaughlin and Wimmer 1999). Decreased radial growth and increased mortality occurred in montane red spruce exposed to acid cloud O3 (Johnson et al. 1988). Sensitivity to damage was enhanced by exposure to mists at acidity levels that occur frequently (DeHayes et al. 1997).
In the southeastern forests, O3 increases effects of soil moisture stress on stem growth of loblolly pine (McLaughlin and Downing 1996). Combined effects of forest growth, natural leaching, and acidic deposition depleted 80% of the base cations from surface soils of a southern pine (Richter and Markewitz 1996). Concentrations of 50-60 ppb O3 were sufficient to cause foliar injury, early needle loss, decreased nutrient availability, reduced carbohydrate production, lower vigor, decreased height/diameter growth, and susceptibility to bark beetles in the San Bernardino Forest of Southern California (Miller and McBride 1999). Diminishing annual O3 concentrations have resulted in improvement in the foliar injury index. The detrimental role of dry N deposition in ponderosa and Jeffrey pine ecosystems is well established (Bytnerowicz and Fenn 1996).
It is clear from the literature that pollutant-induced changes in depth and vigor of root systems, shifts in pool sizes and allocation patterns of carbon, as well as changes in supply of nitrogen and calcium represent important shifts in ecological function occurring in diverse forest types across a large geographic area in North America. These process level changes could be greatly increased in coming decades due to climate change (McLaughlin and Percy 1999).
Air pollution has been linked with declining forest health in a number of regions like the Kola Peninsula (Tikkanen and Niemelä 1995) and Eastern Europe (Innes and Oleksyn 2000). Evidence of O3 injury to trees exists in Greece, Italy (Busotti and Ferretti 1998), Switzerland, and Spain (Skelly et al. 1999). Elsewhere, O3 concentrations may be impacting tree health, and acidification has caused the disruption of ecosystem functions in more sensitive ecosystems (van Breemen and van Dijk 1998). Very high levels of NH3 deposition occur in some areas of the Netherlands and elsewhere (van der Eerden et al. 1998). Acidification problems are well documented in northern countries (Innes and Oleksyn 2000). Critical loads for nitrogen (8 kg ha yr-1) were exceeded at 92% of Level 11 monitoring plots in Europe. When effects on trees are calculated (14 kg ha yr-1 pine; 20 kg ha yr-1 spruce), critical loads are exceeded at 45% of plots. Nitrogen abundance has disrupted more sensitive components, resulting in the loss of species from particular areas (UNECE and EC 2002).
Earlier, Spiecker (1999) found accelerated growth trends in parts of northern Europe, most of central Europe, and in some parts of southern Europe. Decreasing growth trends were found in cases where exposure to pollutants (O3, SO2, N) or exceptional climatic conditions occurred. Subsequent analysis has attributed this increased height growth (1960-1990) to excess nitrogen deposition. Increasing CO2 and climate change are expected to become more important over time as nitrogen effects diminish (European Forest Institute 2002).
One-third of Europe is covered by forest. Despite concerns (Ferretti et al. 1999), annual crown condition survey remains the main tool in the overview on European forest condition. For 1986-1996, deterioration in forest condition was reported (Muller-Edzards et al. 1997), and analysis was unable to explain it through site factors. Although atmospheric concentrations of sulfur and nitrogen were correlated with increased foliar S and N contents (Innes 1995), it was difficult to draw relationships with crown condition. In the most recent (2001) assessment, 22.4% of 132,000 trees assessed were classified as moderately or severely defoliated (UNECE and EC 2002). Temporal development for continuously monitored trees (1989-2001) indicated that (excepting holm oak) mean defoliation of tree species increased in 2001. UNECE and ECE (2002) cite precipitation amount, insects, and fungi as well as air pollution as the most important factors influencing trends. At plots where critical loads are exceeded, increased vulnerability to drought, stress, frost, pests, and diseases is to be expected and diversity of ground vegetation is threatened. In fact, air pollution is considered the most important anthropogenic factor affecting forests in central and eastern Europe. Forest dieback from air pollution occurred over 2.8 million ha of the 39 million ha of forest in this region (EC/PHARE 1999).
Air pollutants have caused changes in tree condition, tree physiology, and biogeochemical cycling; lowered tree resistance to insects and disease; and affected function of diverse forest types in North America. In Europe, critical loads for soils are exceeded, and critical levels for O3 are exceeded at most monitoring plots. Over 22% of trees assessed are still moderately to severely defoliated and acid deposition threatens understorey vegetation.
Yet, we have not always been successful (Table 1) at relating forest health to air pollution. Percy (2002) has concluded that the outcome achieved was dependent on the paradigm within which monitoring systems or studies were working. We must remember that stress responses are non-monotonic. Effects have been shown to cascade from gene-expression all the way through to ecosystem (NPP) levels, (Karnosky et al. 2002). Approaches used to assess forest health must be re-examined if we are to improve detection of future change and elucidation of the roles of natural and anthropogenic stressors.
Table 1. Retrospective analysis of degree of success in documenting the role of air pollution as an important factor in forest health
Network-level monitoring is succeeded by process-oriented research across spatial and temporal scales of stressors
Systematic monitoring is disconnected from process-oriented research
Appropriate indicators of ecosystem function are investigated at an intensity/extent appropriate to stressors and ecosystem
Systematic monitoring is not stratified on stressor gradients
Investigations in essential processes/cycles are integrated with investigations in ecosystem resilience through pests, genetics, succession
System protocols are not developed for single agents
There is continuity of investigation: i.e., Miller and McBride (1999)
Endpoints measured are inappropriate or unresponsive to stressors
Hierarchical nature of forest response is not recognized
Dominant role in predisposition is not recognized
Reproduced from Percy (2002)
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