The Effects of Drought on Amazonian Rain Forests Effects On... · The functioning of Amazonian rain...

429

Amazonia and Global ChangeGeophysical Monograph Series 186Copyright 2009 by the American Geophysical Union.10.1029/2009GM000882

The Effects of Drought on Amazonian Rain Forests

P. Meir,1 P. M. Brando,2,3,4 D. Nepstad,5 S. Vasconcelos,6 A. C. L. Costa,7 E. Davidson,4 S. Almeida,8 R. A. Fisher,9 E. D. Sotta,10 D. Zarin,2 and G. Cardinot11

1School of Geosciences, University of Edinburgh, Edinburgh, UK.

2Department of Botany, University of Florida, Gainesville, Flor-ida, USA.

3Instituto de Pesquisa Ambiental da Amazônia, Belém, Brazil.

The functioning of Amazonian rain forest ecosystems during drought has become a scientific focal point because of associated risks to forest integrity and climate. We review current understanding of drought impacts on Amazon rain forests by summarising the results from two throughfall exclusion (TFE) experiments in old-growth rain forests at Caxiuanã and Tapajós National Forest Reserves, and an irrigation experiment in secondary forest, near Castanhal, Brazil. Soil physical properties strongly influenced drought impacts at each site. Over years 1 to 3 of soil moisture reduction, leaf area index declined by 20–30% at the TFE sites. Leaf physiology and tree mortality results suggested some species-based differences in drought resistance. Mortality was initially resistant to drought but increased after 3 years at Tapajós to 9%, followed by a decline. Transpiration and gross primary production were reduced under TFE at Caxiuanã by 30–40% and 12–13%, respectively, and the maximum fire risk at Tapajós increased from 0.27 to 0.47. Drought reduced soil CO2 emissions by more than 20% at Caxiuanã and Castanhal but not at Tapajós, where N2O and CH4 emissions declined. Overall, the results indicate short-term resistance to drought with reduced productivity, but that increased mortality is likely under substantial, multiyear, reductions in rainfall. These data sets from field-scale experimental manipulations uniquely complement existing observations from Amazonia and will become increasingly powerful if the experiments are extended. Estimating the long-term (decadal-scale) impacts of continued drought on Amazonian forests will also require integrated models to couple changes in vegetation, climate, land management, and fire risk.

4Woods Hole Research Center, Falmouth, Massachusetts, USA.5Gordon and Betty Moore Foundation, Palo Alto, California,

USA.6EMBRAPA-Amazônia Oriental, Belém, Brazil.7Centro de Geociências, Universidade Federal do Pará, Belém,

Brazil.8Museu Paraense Emilio Goeldi, Belém, Brazil.9Los Alamos National Laboratory, Los Alamos, New Mexico,

USA.10EMBRAPA-Amapa, Macapa, Brazil.11Instituto de Pesquisa Ambiental da Amazônia, Brazil.

430 EFFECTS OF DROUGHT ON AMAZONIAN RAIN FORESTS

1. INTRODUCTION

Over the last decade, the possible impacts of drought have become touchstone issues for environmental science and governance in Amazonia. The geographical size, biophysical properties, and species diversity of Amazonian forests have led to analysis of their role as providers of key environmen-tal services at all scales. Increased moisture limitation in the region is likely this century, and some of the services pro-vided by forests, including the storage and sequestration of carbon, evapotranspiration, and the maintenance of species diversity are potentially at risk. Human natural resource use intersects with these roles because of rapid land use change and tends to magnify the likelihoods of drought and forest degradation through fire.

There is uncertainty with respect to all of these outcomes, governed by two core questions: (1) How likely, widespread, and severe is future drought? (2) What is the likely impact of drought on forest ecosystem properties? In this chapter, we address the latter question considering alterations to the carbon balance, transpiration, tree mortality, species-based differences in drought responses, and vulnerability to fire. Evidence of the response by forests to drought is often based on observation of forest processes during natural droughts. However, unusually for any region, three large-scale soil moisture manipulation experiments have been implemented in Amazonia during the Large-Scale Biosphere-Atmosphere (LBA) Experiment in Amazonia. These experiments are pro-viding new tests of the modeled response by rain forest eco-systems to levels of drought that are beyond the bounds of recent climatic variation, but in line with some future cli-mate scenarios. A comparison and analysis of the first re-sults emerging from these experiments forms the core of this chapter and is used to summarize a view of the risks to, and responses by, Amazonian rain forests experiencing drought during the twenty-first century.

2. BACkGROUND

2.1. Future Drought?

Two main lines of evidence suggest that Amazonian drought may become more frequent and more severe during this century. First, episodic drought has been associated with the occurrence of the El Niño–Southern Oscillation (ENSO), caused by the warming of the tropical eastern Pacific Ocean, and more recently in 2005, with abnormal warming of the northern tropical Atlantic Ocean, relative to the south [Cox et al., 2008; Marengo et al., 2008]. Future increases in green-house gas concentrations coupled with reductions in global aerosol emissions may increase the likelihood of 2005-like

drought events [Cox et al., 2008] and possibly also ENSO events [Timmermann, 1999]. More generally with respect to climatic change over the twenty-first century, the multi-model data set (MMD) used in the IPCC Fourth Assessment Report projected an annual mean warming in Amazonia of 1.8°–5.1°C over this century, with rainfall reductions in parts of central and eastern Amazonia likely as a result, especially in the dry season [IPCC Working Group I, 2007]. In a recent analysis involving 23 MMD models, Malhi et al. [2008] re-ported 20–70% agreement among models in the prediction of substantial dry season reductions in precipitation across Amazonia, with the greatest likelihood of drought in the east of the region.

Second, land use change is likely to exacerbate the effects of climatic warming. Widespread forest conversion to pas-ture and agriculture is expected to reduce rainfall in the re-gion through differential effects on latent and sensible heat transfer [Werth and Avissar, 2002; Chagnon and Bras, 2005; Costa et al., 2007; Nobre et al., 1991] and increased regional atmospheric aerosol loading could cause widespread reduc-tions in precipitation [IPCC Working Group I, 2007]. Over-all, these results imply an increased frequency of extreme events at the seasonal and interannual timescale, and a secu-lar shift to drought at decadal timescales, the strength of any shift strongly influenced by land use change.

2.2. Modeling Drought Impacts

Estimates of the impact of drought on Amazonian forests have historically been made with limited field data, and this has contributed to uncertainty in estimates of future global atmospheric CO2 concentrations [Meir et al., 2006; IPCC Working Group I, 2007; Huntingford et al., 2009]. Recent analysis of the bioclimatic distribution of current natural Amazonian vegetation and the predictive output from 19 global circulation models (GCMs) suggests that twenty-first century climate change is most likely to lead to drier condi-tions more appropriate for seasonal forest in eastern Ama-zonia [Malhi et al., 2009], although edaphic conditions may in reality favor a transition to degraded forest or savanna. In contrast, smaller impacts are likely on vegetation in western Amazonia [Malhi et al., 2009]. Driving a vegetation model with a variety of climate scenarios, Sampaio et al. [2007] predicted a similar but more extreme range of forest-to-sa-vanna switches following climatic warming and drying. In both cases, these decadal-century timescale scenarios were based on some form of “equilibrium” vegetation response to drought. In reality, the actual vegetation response will be determined (rapidly or gradually) by multiple ecologi-cal and physical processes, as acknowledged by the same authors.

MEIR ET AL. 431

By contrast, process-based dynamic vegetation models have the structure to capture the relevant ecology, enabling land surface-atmosphere interactions to be modeled on a continuous basis. However, they are computationally ex-pensive and ultimately require observation-based, ecologi-cal parameterization [Prentice and Lloyd, 1998; Meir et al., 2008]. The first predictions of substantial Amazonian die-back in response to warming and drought emerged from a global dynamic vegetation model (DGVM) framework and were made at the decadal-century timescale, with relatively simplistic representations of canopy structure, soil processes, and functional diversity among species [White et al., 1999; Cox et al., 2000]. Large differences in parameterization and mathematical description of core ecosystem processes have since been identified among DGVMs [Dufresne et al., 2002; Meir et al., 2006; Friedlingstein et al., 2006; B. Poulter et al., Managing uncertainty of tropical Amazon dieback, sub-mitted to Global Change Ecology, 2009; D. Galbraith et al., Quantifying the contributions of different environmental factors to predictions of Amazonian rain forest dieback in three dynamic global vegetation models (DGVMs), submit-ted to Global Change Biology, 2009], emphasizing the need for improvements using field-based measurements and ex-perimentation. Nondynamic, but process-based, vegetation and biogeochemical models have also been used to advance the representation of Amazonian forest functioning [Lloyd et al., 1995; Williams et al., 1998; Potter et al., 2004], al-though differences in process description have been ap-parent [Tian et al., 1998; Prentice and Lloyd, 1998; Foley et al., 2002; Zeng et al., 2005; Meir et al., 2008]. Only re-cently have detailed validations been made of the modeled response to drought and with some success over seasonal to interannual timescales [Fisher et al., 2007; Baker et al., 2008].

The validation of longer-term vegetation dynamics is more difficult because it requires data sets at the scale of years to decades, representing multiple ecological processes. The multiyear soil moisture manipulation experiments de-scribed here provide unique insight into some of the relevant physiological and ecological processes. Also, they indicate the potential first steps in quantifying natural lags in veg-etation change as a function of tree species’ drought resist-ance and regeneration characteristics. However, vegetation change under climatic drying is influenced strongly by fire and deforestation. The area of burnt understory forest during the ENSO of 1998 (2.9 × 106 ha) was more than 10 times greater than during an average rainfall year and twice the area of annual deforestation [Nepstad et al., 1999; Alencar et al., 2006], while a burnt area up to 2800 km2 was attrib-uted to fire leakage alone in the 2005 drought [Aragão et al., 2007]. Fire is more probable in forest that has burned

previously [Cochrane, 2003], and the fragmentation effects of deforestation tend to increase fire risk strongly [Uhl and Kauffman, 1990; Cochrane and Laurance, 2002]. Deforest-ation has repeatedly been shown to affect regional climate, though the effects vary with scale [Werth and Avissar, 2002; Chagnon and Bras, 2005]. At large scale, deforestation scenarios result in modeled reductions in precipitation and relative humidity, and increases in temperature [e.g., Hoff-mann et al., 2003; Costa et al., 2007]. Strong increases in fire risk (20–120%) have thus been associated with scenarios of partial and complete deforestation [Cardoso et al., 2003]. While we need to understand the responses to drought by Amazonian rain forest, the interactions between drought and increased fire risk must also be considered, as fire may well be the agent that, in the context of secular or episodic cli-matic drying, triggers the switch from a forest to a savanna [Hutyra et al., 2005; Aragão et al., 2007].

2.3. Basin-Wide Observations of the Forest Carbon Balance During Drought

Inversion studies have provided the largest scale of obser-vation-based information on drought effects on Amazonia. Despite the limitations of this method [see Houghton et al., this volume], a net emission of CO2 from the region of up to 1.5 Pg C a−1 has been reported for dry and warm periods during strong ENSOs [Bousquet et al., 2000, Rödenbeck et al., 2003, Zeng et al., 2005, see also Houghton et al., this volume], although it is unclear if these higher emissions are principally the result of changes in ecosystem carbon cy-cling or increased fire occurrence [Langenfelds et al., 2002; Meir et al., 2008]. Results from biogeochemical modeling of the forest carbon cycle have been consistent with the ob-servation of net regional CO2 emissions during ENSO [e.g., Tian et al., 1998; Foley et al., 2002; Zeng et al., 2005], but the mechanisms underpinning these modeled results have tended to overemphasize the role of the temperature sensi-tivity of soil respiration during drought [Meir et al., 2008]. Phillips et al. [2009] recently reported basin-wide long-term observations of Amazonian forest productivity, estimating an overall negative impact of 1.2–1.6 Pg C in old growth forests during the 2005 drought, which was driven by lower growth and (spatially patchy) increases in tree mortality. This outcome is consistent with the notion that gross pri-mary production (GPP) declines under drought, but without measurements of concurrent soil processes (e.g., CO2 efflux from soil), it does not resolve the question of the short-term impact of the 2005 drought on net ecosystem productivity (NEP). [NEP is the difference between carbon that is photo-synthesised and respired by an ecosystem. NEP = GPP − Reco, (Reco = total ecosystem respiration).]


In summary, analysis of basin-wide atmospheric and tree growth data has highlighted the importance of understand-ing NEP at different timescales, and identified large changes in possible ecosystem functioning during interannual-scale drought, including reductions in aboveground productivity, increases in mortality in response to drought, and differential tree species’ survival [cf. Engelbrecht et al., 2007]. These data will be invaluable to validate estimates of tree growth from DGVMs or finer-scale vegetation models, but alone, they do not pinpoint the mechanisms determining large-scale CO2 exchange observed during intermittent severe drought [e.g., Rödenbeck et al., 2003]. In particular, CO2 emission processes have been hard to pin down, first because respira-tion terms are often poorly quantified [Meir et al., 2008], and second, because the increases in fire occurrence identi-fied during ENSO or the 2005 drought [Nepstad et al., 1999; Aragão et al., 2007] are hard to quantify in terms of CO2 emissions [van der Werf et al., 2004].

2.4. Observations at the Stand Scale

Some of the first eddy covariance measurements in Ama-zonia, quantifying the forest water and carbon cycles at the scale of a few square kilometers, suggested little seasonality in rain forest carbon exchange capacity [Grace et al., 1995]. However, although not universal, seasonal drought effects on NEP have since been observed at some sites, with observed changes in gross photosynthesis, respiration, and transpira-tion across sites differing in seasonality yielding new in-sights into the functional basis of the response [Malhi et al., 1998; Carswell et al., 2002; Saleska et al., 2003; Vourlitis et al., 2005; see da Rocha et al., this volume; Saleska et al., this volume; da Rocha et al., 2009].

Uniquely, experimental field manipulation of soil moisture enables the separation of the effects of otherwise naturally covarying environmental drivers (i.e., edaphic and atmo-spheric variables) of the processes determining NEP and has the potential to provide powerful insight into process-level responses to short-term and extended drought. Three such field experiments have been performed within LBA, all in drought-threatened eastern Amazonia. Strong soil moisture deficit has been imposed over 1 ha of old-growth forest at two experimental sites, in the National Forest Reserves of Caxiuanã (near Portel, State of Pará) and Tapajós (near Santarém, State of Pará) [Nepstad et al., 2002; Fisher et al., 2007; Meir et al., 2008], and at a third site, dry-season moisture stress has been reduced though the irrigation of replicated 0.04-ha plots in secondary regrowth forest near Castanhal, State of Pará [Vasconcelos et al., 2004].

The remainder of this chapter is concerned primarily with how the results of these experiments inform our understand-

ing of drought impacts on Amazonian rain forest, mostly in relation to the cycling of carbon and water, but also includ-ing the emission of other trace gases such as CH4, NO, N2O, and isoprene. Where possible, we make a first comparison of the still-emerging experimental results. We ask three questions: (1) What have we discovered about the impact of drought at seasonal to interannual timescales? (2) What have we discovered about the impact of drought at multiyear to decadal timescales? (3) How does the combined impact of fire and drought influence the risk of widespread loss of forest in favor of smaller-stature vegetation types, such as savanna or degraded forest?

3. SOIL MOISTURE MANIPULATION ExPERIMENTS AT CAxIUANã, TAPAjóS, AND CASTANHAL

Species diversity is substantial (>150–200 species ha−1) in the old-growth rain forests at both Caxiuanã and Tapa-jós, while only four tree species (=70% of stems) dominate the regrowth forest at Castanhal. Standing biomass is much larger at the rain forest sites (approximately 300 t ha−1 at Caxiuanã, 240 t ha−1 at Tapajós, and 50 t ha−1 at Castanhal) [Baker et al., 2004; Vasconcelos et al., 2004; Brando et al., 2008]. Rain forest canopy heights are similar at 30–40 m, while the height of the secondary forest at Castanhal is ap-proximately 5 m [Coelho et al., 2004]. The soils at Caxiuanã and Tapajós are highly weathered Oxisols, and at Castan-hal, the soil is a dystrophic yellow Latosol. At Caxiuanã, the soil composition is a sandy loam (70–83% sand) with the water table at 15–20 m depth; at Tapajós, the composi-tion is more clay-rich and the soil profile is much deeper (60–80% clay; >100 m in depth); and at Castanhal, the soil is shallow and concretionary, with a high sand content (20% clay, 74% sand). Annual precipitation at Caxiuanã, Tapajós, and Castanhal is approximately 2300, 2000, and 2500 mm a−1, respectively. Further site descriptions are provided else-where [Davidson, 1992; Nepstad et al., 2002; Ruivo et al., 2003; Fisher et al., 2007; Vasconcelos et al., 2004].

The method of physically excluding rainfall that penetrates the canopy (“throughfall exclusion” (TFE)) was replicated at Caxiuanã and Tapajós using approximately six thousand 4.5 m2 plastic panels and guttering placed at 2 m above the ground. The infrastructure removed approximately 50% of incoming precipitation [Nepstad et al., 2002; Fisher et al., 2007] (Table 1) and was installed at the beginning of 2000 at Tapajós and 2002 at Caxiuanã. Each experiment comprised 1 ha of TFE forest and 1 ha of undisturbed (nonmanipulated) “control” forest. The large scale of the manipulation was nec-essary because of substantial lateral extension of the surface roots of large trees. Treatment replication at both sites was limited by financial resources, but pretreatment calibration

MEIR ET AL. 433

measurements were made in all plots to enable replication over time [Davidson et al., 2004], and the method follows the design of other unreplicated large-scale ecosystem ma-nipulation experiments [e.g., Likens et al., 1970], whose strength is acknowledged, especially where large treatment effects are expected [Hurlbert, 2004]. The perimeter of the TFE plots was trenched to 1–2 m depth to prevent the hori-zontal ingress of water from adjacent normally watered soil, and the control plot perimeters were also trenched to avoid confounding treatment effects. Soil and plant measurements were made more than 20 m inside the perimeter of each plot to further eliminate confounding treatment effects. Manage-ment of the experiments was similar, except that at Tapajós, the panels were removed during the peak dry season, while at Caxiuanã, this procedure was not followed because of the prevailing late dry season storm risk [Carswell et al., 2002]. Litterfall was manually returned to the soil, where it fell on the TFE paneling. Full canopy access was provided using 40 m towers in all plots, also enabling the installation of on-site automatic weather stations. Principal access to the soil was provided by four soil shafts per plot, excavated to a maxi-mum depth of 10 m at Caxiuanã and 14 m at Tapajós.

The irrigation experiment at Castanhal was designed to remove dry season moisture stress and began in 1999 [Vas-concelos et al., 2004]. Plot size was 20 m × 20 m (0.04 ha), and four replicate plots were used to contrast irrigated and undisturbed vegetation; adjacent plots were placed 10 m or more apart and a nested 10 m × 10 m plot in each 0.04 ha main plot was used for measurements. An additional treat-ment removing litter was also implemented at Castanhal, but is not discussed here. Ground-level irrigation was provided using perforated microtape installed at 2-m spacings; water was applied at a rate of approximately 5 mm d−1 for 30 min during the dry season. The amount of irrigation was selected to approximately replace daily evapotranspiration estimated regionally (660–790 mm) [Jipp et al., 1998, Vasconcelos et al., 2004]. Surface soil moisture availability was measured

in all plots, and the relatively small stature of the vegetation allowed lower-canopy access for leaf water potential mea-surements [Fortini et al., 2003].

4. SEASONAL TO INTERANNUAL DROUGHT IMPACTS

4.1. Soil Moisture and Its Supply to Plants

The change in soil moisture, in relation to adjacent un-disturbed (control) forest, was the main experimentally manipulated parameter in each experiment. The TFE infra-structure at Caxiuanã and Tapajós resulted in reductions in plant available water (PAW) of 80–200 mm in the top 3 m of soil (Table 1 and Figure 1a) [Fisher et al., 2007, 2008; Brando et al., 2008], while the irrigation at Castanhal nearly completely removed the dry season moisture constraints [Vasconcelos et al., 2004]. A strong seasonality was evident in PAW at the TFE experiments, and the rate of soil moisture drawdown was larger over the first 1–3 years at Caxiuanã than at Tapajós.

The impact on PAW of manipulating precipitation input to the soil was strongly modified by the soil properties at each site. At Caxiuanã, the sandy-loam composition created a relatively high moisture holding capacity per unit volume [Carswell et al., 2002; Fisher et al., 2008], while the deep clay-rich soil at Tapajós potentially held substantial water reserves mainly because of its exceptional soil volume [Nep-stad et al., 2002; Belk et al., 2007]. Detailed measurements of soil hydraulic properties [e.g., Tomasella and Hodnett, 1997] remain rare in Amazonia, but variation in the key soil parameters determining PAW, soil water potential, hydrau-lic conductivity, and moisture volume content, can cause large and basin-wide differences in the supply of water to plants under moisture stress [Fisher et al., 2008]. Analo-gously, the shallow concretionary soil structure at Castanhal likely contributes to moisture limitation of root and micro-

Table 1. Total Incoming Precipitation for Each Year of the Drought Treatment by Dry (Panels Off) and Wet Season (Panels On)a

Precipitation, mm

Year Period of Exclusion Total Incoming Panels OffPanels On

(Throughfall Excluded)

2000 1 Feb to 8 july 2517 830 1687 (844)2001 1 jan to 31 july 1882 171 1711 (856)2002 1 jan to 31 july 1958 292 1665 (833)2003 21 jan to 14 Aug 1690 394 1295 (648)

Totals 8047 1687 6358 (3179)aApproximately 50% of incoming precipitation was excluded when the panels were on; those

volumes are presented in parentheses.


Figure 1. (a) Plant available water (PAW) over wet and dry seasons in the control (filled symbols) and throughfall ex-clusion (TFE) plots (open symbols) for both experiments Tapajós (circles) and Caxiuanã (triangles). The experiment in Caxiuanã began in 2001, and in Tapajós, it started in 1999 (pretreatment). PAW is the water available to plants between field capacity and a soil moisture potential of −1500 kPa. The arrow shows when TFE infrastructure was installed at each site. The data points are derived from 48 fortnightly measurements made per soil access shaft, four access shafts per plot [Brando et al., 2008; Fisher et al., 2007]. (b) Monthly natural rainfall, soil water balance, and annual soil water deficit at Caxiuanã, 1999–2004. Maximum PAW at Caxuiana is ~400 mm water (a) [Fisher et al., 2008]; under normal rainfall, soil moisture supply is sufficient to support transpiration, but under strong drought, moisture limitation is likely.

(a)

(b)

MEIR ET AL. 435

bial activity during the natural dry season [Vasconcelos et al., 2004]. More soil hydraulics data from across the region are required to make satisfactory PAW calculations for the modeling of vegetation activity.

The supply of moisture from soil to leaf is also affected by site differences in rooting properties. Rooting depth can be substantial, enabling increased access to soil water: roots have been detected at depths >14 m at Tapajós and at Cax-iuanã, to the approximate maximum depth of excavation, 9 m [Nepstad et al., 2004; Fisher et al., 2007]. As well as enabling fuller exploration of the soil profile, deep rooting can also facilitate hydraulic redistribution, thus further help-ing to maintain rhizosphere moisture availability and plant function under reduced rainfall, as observed at Tapajós in some instances [Oliveira et al., 2005]. Initial data have not identified major changes in root dynamics at Tapajós in the TFE treatment [Brando et al., 2008]. However, responses in the modes of root growth at Caxiuanã are partially con-sistent with theory [Joslin et al., 2000; Schymanski et al., 2008]. As the drought progressed at Caxiuanã, surface roots (0–30 cm) tended to increase in length per unit mass, thus increasing the explored soil volume [Metcalfe et al., 2008], but it is less clear how plastic root growth responses were to changes in the vertical distribution of soil moisture availabil-ity within the soil profile, and answering this question will have implications for modeling resilience to drought. Spe-cies variation in these root properties differentially affects water acquisition among species and between sites, and vari-ation in resistance to drought was also observed in the foli-age. Leaves generally experience maximum daily moisture limitation at high atmospheric moisture deficit, soon after midday. Early afternoon measurements of the minimum leaf water potential tolerated by tree species at Caxiuanã, Tapa-jós, and Castanhal demonstrated differences among species (Table 2), although the minimum value measured in both TFE plots was similar (−3.2 and −2.7 MPa, respectively), indicating a possible maximum tolerance to moisture limita-tion in eastern Amazonian rain forest trees. Stem hydraulic conductance did not appear to be the major constraint to leaf water supply at low soil PAW, but species-based differences in this parameter were also observed at Caxiuanã [Fisher et al., 2006] potentially further influencing species-based dif-ferences in resistance to soil moisture limitation [cf. Franks et al., 2007].

In summary, substantial differences in soil properties at each experimental site strongly influenced the storage of wa-ter in soil and its supply to plant roots under drought. The TFE results from Tapajós and Caxiuanã also highlighted mechanisms conferring tolerance to drought in terms of root, stem, and leaf hydraulic properties. Variation among tree species in these responses to drought, including addi-

tional evidence of direct uptake of dew in two species at the Tapajós experiment (G. Cardinot et al., manuscript in re-view, 2009) suggest likely differences in survivorship under drought at multiyear or decadal time scales.

4.2. Canopy Structure and Productivity

Although some savanna tree species have a deciduous or brevi-deciduous phenology [Furley et al., 1992], this strat-egy is relatively unusual in rain forests, where leaf area index (LAI, m2 leaf area per m2 unit ground area) is maintained un-der normal climatic variation. There is some ground-based evidence for dry-season increases in LAI [Carswell et al., 2002] and albedo [Culf et al., 1995], but irrespective of how this may affect forest functioning, it seems that PAW can often be maintained by forest trees through extensive and sometimes exceptionally deep rooting systems [Nepstad et al., 1994; Bruno et al., 2006]. However, under moisture constraints in excess of normal climatic variation, we have limited understanding of the limits to moisture access by for-est trees. The TFE experiments thus provide a direct way of determining the thresholds in PAW that may lead to changes in canopy productivity, LAI, and fire vulnerability.

The LAI at both sites was resistant to strong artificial soil moisture deficit for about a year following installation of the TFE infrastructure, remaining at 5–6 m2 m−2. After 12 months, and a reduction in PAW of about 150–200 mm, LAI declined to 70–80% of the control (and original) values at each site (Figure 2), and this reduction was maintained subsequently following three more years of TFE (Brando et al., 2008, D. B. Metcalfe et al., Impacts of experimen-tally imposed drought on leaf respiration and morphology

Table 2. Minimum Leaf Water Potential (Min ψl) Measurements for Tree Species at the Caxiuanã and Tapajós TFE Experiments and the Castanhal Irrigation Experimenta

Site Species Min ψlb, MPa

T Aparisthmium cordatum −1.5T Astrocaryum gynacanthum −1.5T Coussarea racemosa −2.7T Poecilanthe effusa −2.4C Licaria ameniaca −2.2C Hirtela bicornis −2.1C Lecythis confertiflora −2.9C Swartzia racemosa −3.2Cs Miconia ciliate −3.0

aAbbreviations are C, Caxiuanã; T, Tapajós; Cs, Castanhal.bValues for T and C are taken from trees in the TFE plots, and

values for Cs are from an understorey species (data from Fisher et al. [2006], Fortini et al. [2003], and G. Cardinot et al., manuscript in review, 2009).


Figure 2. Variation in leaf area index (LAI) during the TFE experiments at Caxiuanã and Tapajós. LAI is expressed as the quotient of (TFE plot LAI) / (Control plot LAI). LAI in undisturbed forest at both sites is 5–6 m2 m−2. The arrow denotes when the TFE infrastructure was installed at each site. (LAI was measured using a Li2000 (Licor, USA) at fixed 100 points [Brando et al., 2008; Fisher et al., 2007; Metcalfe et al., submitted manuscript, 2009]).

in an Amazon rain forest, submitted to Functional Ecology, 2009). However, litter production patterns differed among sites, perhaps partially reflecting differences in tree commu-nity responses to moisture limitation. At Caxiuanã, litterfall in the TFE plot declined within the first 12 months following installation of the TFE and remained lower than undisturbed forest over the following dry seasons. At Tapajós in the TFE plot, and at Castanhal under irrigation, litterfall did not change significantly and tracked the control forest litter flux rates closely in the first 2–3 years of rainfall manipulation (Figure 3). The reduction in LAI and subsequent litterfall at Caxiuanã is consistent with reduced leaf regrowth, while the maintenance of litterfall at Tapajós and Castanhal suggests a stronger limitation on leaf replacement. However, observed increases in leaf mass per unit area during experimental drought also contributed substantially to changes in LAI at Caxiuanã and Tapajós [cf. Wright et al., 2006; Metcalfe et al., submitted manuscript, 2009; Tohver et al., unpublished data, 2007]. Finally, changes in leaf turnover rates, and the temporary impacts of mortality events after three years or more (data not shown) may help explain litterfall patterns over the longer term at Tapajós and Castanhal [Brando et al., 2008; Vasconcelos et al., 2008].

In summary, initial resistance to change in LAI during the first 12 months of soil drought at both TFE experiments was followed by substantial reductions in LAI of 20–30% over the following 2 years, and this was maintained subsequently. Dry season litterfall declined at Caxiuanã relative to undis-turbed forest, but artificial droughting and irrigation had only small effects on litterfall at Tapajós and Castanhal. Drought impacts on forest canopy physiology and structure have also been examined using remote sensing data products [Asner et al., 2004; Saleska et al., 2007; Huete et al., 2008, see also Saleska et al., this volume]. It remains unclear if these data can be used to reliably quantify changes in productivity during drought, but the TFE experiments provide a way of specifying models to test such estimates. The consequences of reduced LAI for alterations in gas exchange capacity and fire vulnerability are considered below.

4.3. Trace Gas Emissions From Soil

The flux of CO2 from soil (“soil respiration,” Rs), com-prises the largest single respiratory flux in the terrestrial carbon cycle and derives from the combined respiration of heterotrophic (microbial, faunal) and autotrophic (root)

MEIR ET AL. 437

components of the soil biological community [Trumbore, 2006; see also Trumbore and de Camargo, this volume]. Rs has been reported for all three experiments, and additional measurements of NO, N2O, and CH4 have also been made at Tapajós and Castanhal.

Consistent with biophysical expectation [Howard, 1979; Meir et al., 2008], seasonal declines in Rs under reduced soil moisture have been observed widely in Latin American rain forests [Davidson et al., 2000; Schwendenmann et al., 2003]. While the experimental manipulation of soil moisture lowered Rs strongly during the first 2 years at Caxiuanã (by >20%, equivalent to >2 t C ha−1 a−1; Figure 4) [Sotta et al., 2007], the response of Rs to temperature was small and non-

significant [Sotta et al., 2007]. After 3–4 years, dry season Rs in the Caxiuanã TFE plot remained lower than for soil in undroughted forest, although overall between-plot dif-ferences in Rs were smaller, perhaps because of increased wet-season root respiration rates in the TFE plot [Metcalfe et al., 2007]. At Castanhal, the difference in Rs between ir-rigated and undisturbed soil reached maxima during the dry seasons, and annual Rs in irrigated plots was 13–27% larger than on undisturbed (drier) plots [Vasconcelos et al., 2004]. In contrast, at Tapajós, although the gross fluxes were not unusual in magnitude, Rs was similar between the TFE and control plots throughout the experiment, even after 5 years of TFE treatment [Davidson et al., 2008]. This outcome was

Figure 3. Seasonal litterfall (±SE) at Caxiuanã, Tapajós, and Castanhal from 1999 to 2004. The arrows denote when the TFE or irrigation treatments were installed. The data are derived from 20 litter traps (1 m2) per plot; litterfall was collected fortnightly or monthly at each site [Brando et al., 2008; Vasconcelos et al., 2008; Meir et al., manuscript in preparation, 2009].


surprising given the seasonal variation in PAW at Tapajós, and the close tracking of Rs with seasonal PAW in the TFE plot at Caxiuanã (Figures 1 and 4). The similarity in Rs be-tween TFE and control plots at Tapajós is consistent with the similarity in total litterfall and the radiocarbon-derived ages of respired carbon from both plots. This has led to the suggestion that the unusually large soil and rooting depths at Tapajós explains the unexpected maintenance of normal Rs fluxes under drought, although increased root mortality and decomposition in the TFE plot may also have been responsi-ble [Brando et al., 2008; Davidson et al., 2008]. An additional possible explanation is that the role of hydraulic redistribu-tion has been underestimated. For example, if nighttime re-covery of rhizosphere moisture content occurred at Tapajós,

but was not detected using the time domain reflectometry instruments installed there [Nepstad et al., 2002; Oliveira et al., 2005], then autotrophic and heterotrophic respiration may have continued in the roots and rhizosphere, respec-tively, maintaining relatively high Rs even at the low PAW measured in the TFE plot, although how (or if) Rs and GPP were differentially affected by such levels of rhizosphere moisture availability remains unexplored at Tapajós.

Emissions of NO, N2O, and CH4 were also measured at Tapajós and Castanhal. Moisture limitation effects on each followed biophysical expectation [Forster et al., 2007], al-though NO emissions were more resistant to change than expected. At the Tapajós TFE experiment, N2O emissions declined, and CH4 consumption increased at low soil mois-

Figure 4. Soil respiration (CO2 efflux from soil, Rs) at Caxiuanã, Tapajós, and Castanhal. TFE or irrigated plots (gray triangles) are plotted against control plots (black filled circles) at each site. Measurements were made by infrared gas analysis, using permanently installed collars. Data are presented as monthly means ±SE (n = 20 at Caxiuanã and Tapajós, n = 12 at Castanhal), with the experimental soil moisture manipulations beginning at month 0. Inset: the relationship between soil moisture potential and Rs from Caxiuanã. Data are from Vasconcelos et al. [2004], Sotta et al. [2007], and Davidson et al. [2008]. Inset from Sotta et al. [2007], reprinted with permission from Wiley-Blackwell.

MEIR ET AL. 439

ture availability. Consistent with this, irrigation at Castanhal had the reverse effects on fluxes of both trace gases; the thin concretionary soils and secondary regrowth history of this site were associated with lower overall fluxes [Vasconcelos et al., 2004]. Emissions of NO were not altered by irrigation at Castanhal, and they did not increase substantially at the Tapajós TFE until soil moisture availability was very low, a result also partially attributed to the soil texture at this site [Vasconcelos et al., 2004; Davidson et al., 2008]. Follow-ing permanent removal of the TFE treatment after 5 years at Tapajós, Davidson et al. [2008] reported a return of all trace gas soil fluxes to pre-TFE treatment levels and argued that the observed effects of the drought treatment at Tapa-jós most likely reflected changes in soil aeration rather than substrate supply.

In summary, experimental manipulation quantified the strong influence of soil moisture on Rs in two of the three experiments (at Caxiuanã and Castanhal). These results demonstrated the primary importance of moisture limitation over temperature on respiration processes during drought, consistent with observations elsewhere in the region and, contradicting earlier, widely employed, modeling assump-tions [Tian et al., 1998; Zeng et al., 2005; Peylin et al., 2005]. The maintenance of relatively high Rs at low soil moisture content observed in the TFE plot at Tapajós, possibly ex-plained by an exceptionally deep soil profile, serves to high-light the existence of spatial variation in drought responses across the basin, although other drought-related trace gas emissions of NO, N2O, and CH4 responded to reduced soil moisture availability as expected, with temporary reductions in CH4 and N20 emission and increases in NO production under severe moisture limitation.

4.4. Fire Risk

One of the most obvious ways in which drought affects the flammability of tropical forests is through temporary changes in the understory microclimate: drier and warmer conditions increase the risk of fire [Alencar et al., 2004, 2006]. But less obvious are the indirect and lasting effects of drought on forest flammability through reduced PAW [Nep-stad et al., 1999, 2001, 2004]. As PAW reaches deficits large enough to induce leaf shedding, solar radiation penetrating the forest canopy increases and leads to higher air tempera-tures near the forest floor [Uhl and Kauffman, 1990]. This process speeds the rate at which the drying of fine fuel (e.g., leaves and small twigs) occurs, one of the best proxies of forest flammability [Hoffmann et al., 2003].

The observed LAI reductions in both TFE experiments were sufficient to increase forest flammability in the TFE plots. Based on an LAI and precipitation-driven model

developed from experimental fires conducted in the vicinity of the Tapajós TFE experiment [Ray et al., 2005], we calcu-lated daily probabilities of a fire spreading in both control and TFE plots for the Tapajós experiment, from july 2000 to December 2004 [equation (1), Figure 5]. We ran two simu-lations: one in which both precipitation and LAI varied and another in which precipitation was set to zero while LAI was not constrained.

5.35 0.3*cwp 0.131*CH 0.36*LAI11

1 - - -= -+

Pe

, (1)

where cwp is the sum of precipitation in time (t) in days and the cumulative precipitation in time (t − 1) × 0.5; CH is canopy height (kept constant); and LAI is monthly leaf area index.

During the dry season of 2002, the maximum probability of fire spread (P) in the control forest was 0.27, while in the TFE, plot P was 0.47. This increase in fire susceptibility occurred especially after the pulse in tree mortality, 3 years following installation of the TFE infrastructure. By simulat-ing forest flammability based only on understory microcli-mate, we assumed that fine fuel loads would not limit fire spread, although we note that the increase in coarse woody debris resulting from higher mortality would have increased fire intensity.

Figure 5. Probability of fire spread for the Tapajós TFE (gray) and control (black) plot forests. Dashed black and gray lines represent where weighted precipitation was set to zero, so LAI is driving for-est flammability. Fire spread probability is calculated using equa-tion (1) and data from the TFE experiment [Ray et al., 2005].


In summary, the experimental drought treatment substan-tially increased the susceptibility of the forest to fire, even over the short term (Figure 5), converting it from a state where fire was unlikely to one where fire was probable in the presence of appropriate ignition sources. The risk of fire thus increases markedly under extended drought, and this has implications for long-term carbon storage and emissions [van der Werf et al., 2004]. The decline of LAI in both TFE experiments, and the particularly high mortality at Tapajós, signify that the positive feedbacks between drought and fire may be stronger than previously hypothesized.

5. MEDIUM- AND LONG-TERM DROUGHT IMPACTS: RESPONSES IN PHYSIOLOGY AND MORTALITY

Physiological responses to drought observed over the short term have impacts over the long term, most notably through their negative impacts on growth and mortality. Tree death delivers carbon to the decomposer pool, committing future emissions of CO2 to the atmosphere, and in the absence of replacement through recruitment and regrowth, overall net primary production and transpiration decline, while the risk to fire rises.

McDowell et al. [2008] review proposed plant physiologi-cal responses to drought [Tardieu and Simonneau, 1998] and distinguish between plants under moisture limitation that exert strong stomatal control to maintain leaf water potential above a minimum value (“isohydry”) and those that exert less stomatal control of transpiration, relying on continued water supply from the soil (“anisohydry”). Under strong moisture limitation (and high tension), the transpira-tion stream may cleave (embolize), and if this happens to a large extent, hydraulic failure can occur, leading to plant death [Tyree and Sperry, 1988; West et al., 2008]. Anisohy-dric plants exert minimal stomatal control under moisture limitation and thus risk mortality by hydraulic failure un-der severe drought unless this risk is reduced, for example, through the construction of resistant xylem vessels. By con-trast, isohydric plants reduce the risk of hydraulic failure through stomatal closure, although this risk is not entirely avoided, and other resistance terms in the soil-to-atmosphere hydraulic path may also change [Franks et al., 2007]. Sto-matal closure can lead to high leaf temperatures, to reduced photosynthetic carbon gain, and under extended drought to increased risk of mortality through carbon starvation of me-tabolism (mainly respiration) and/or increased susceptibility to pathogen attack.

Leaf water potential measurements from the Caxiuanã TFE experiment were consistent with trees responding iso-hydrically to drought [Fisher et al., 2006]. This proposed mode of response is further consistent with: (1) declines in

GPP and stem growth rates following TFE treatment [Fisher et al., 2007; Brando et al., 2008]; (2) the initial resistance to mortality observed over the first 2 years of experimen-tal drought at Caxiuanã and Tapajós [Nepstad et al., 2007; A. C. L. Costa et al., manuscript in preparation, 2009]; and (3) pantropical observations of small average increases in mortality following ENSO events over the last three dec-ades (the median increase in mortality following an ENSO event across 45 tropical forest plots was 1.2%) [Meir et al., 2008].

Further consistent with the notion that isohydry may be common in rain forest trees that are not adapted to long-term and severe droughts, resistance in mortality rates to the TFE treatment gave way after 2–3 years to substantial mortality increases [Nepstad et al., 2007; Brando et al., 2008], possi-bly as a result of carbon starvation. The Tapajós TFE experi-ment revealed highly variable background rates of mortality in both the TFE and control plots (1–3%), and this was then followed by exceptionally high mortality (9% in trees with dbh > 10 cm) after 3 years of experimental drought [Nepstad et al., 2007; Brando et al., 2008]. In years 4 and 5 of the experimental drought at Tapajós, mortality declined to just above (the relatively high) background levels, but 1 year after removal of the TFE infrastructure at Tapajós, mortality rose again to 7%, suggesting longer term and possibly species- specific impacts. Indications of a correlation between spe-cies differences in stomatal control and mortality at Tapajós [Ehleringer et al., 2004] further support the notion of spe-cies-specific variation in mortality risk during drought [cf. Fisher et al., 2006] (see Table 2), and increased regrowth during years 4 and 5 by understory species released via mor-tality impacts on the canopy will also have influenced the range of species-based responses at Tapajós [Brando et al., 2008]. The immediate influence of increased mortality on CO2 emissions is likely to be small during drought because of the desiccation constraint on organic matter breakdown [Meir et al., 2008], although the effect of such strong pulses of tree death on ecosystem-level GPP is less clear because mortality reduces LAI while simultaneously increasing radi-ation availability within the canopy. However, over the long term, mortality clearly commits substantial carbon to the at-mosphere through the breakdown of additional necromass. If the high mortality pulse observed at the Tapajós TFE experi-ment after 3 years of drought (5.4 Mg C ha−1) [Brando et al., 2008] occurred widely over the region, it would represent a large net committed emission to the atmosphere. Although mortality at Tapajós declined to 2–4% under the following 2 years of TFE treatment [Brando et al., 2008], the live bio-mass removed during this single mortality event represented up to 8.5 years of aboveground growth under normal rainfall [Nepstad et al., 2007].

MEIR ET AL. 441

The findings from the TFE experiments are consistent with recent observational reports of natural drought effects on Amazonian forest trees. For example, following the severe drought event of 2005, tree growth observations in 55 1-ha plots distributed across Amazonia demonstrated drought- induced reductions in stem growth (especially in larger trees) and significant, but spatially patchy, increases in mortality [Phillips et al., 2009]. The 2005 drought in Amazonia was less severe or prolonged than the soil moisture limitation imposed in the TFE experiments and comprised additional climatic impacts on temperature, atmospheric humidity, and radiation, yet Phillips et al. [2009] estimated an overall re-duction in aboveground growth of 5.3 Mg ha−1, in addition to a substantial increase in mortality-committed carbon emis-sions. The regional-scale spatial variation in mortality ob-served during 2005 was dependent on differences in climate, soil-type, and species, with a tendency for higher mortality in species with lower density wood [Phillips et al., 2009]. Such spatial variation in the response to drought was also evident in a recent pantropical survey of recent ENSO im-pacts on mortality [Meir and Grace, 2005; Meir et al., 2008]. Taken together with evidence of species differences in pho-tosynthesis, growth, and reproduction from the TFE experi-ments [Ehleringer et al., 2004; Fisher et al., 2006; Brando et al., 2006], these results imply initial, but spatially variable, resistance to short-term soil moisture limitation, followed by increased mortality and likely alterations in tree community composition as the severity of drought deepens and extends.

The longer term (multiyear to decadal) effects of drought on NEP are not well constrained by any current data sets, but may contain surprises. As observed in the TFE experi-ments and elsewhere, short-term reductions in GPP and Rs are likely, and increased mortality and fire incidence will in-crease losses of carbon to the atmosphere. However, recent observations of significant medium-term (5 year) increases in leaf respiration at the Caxiuanã TFE experiment [Meir et al., 2008; Metcalfe et al., submitted manuscript, 2009] sug-gest unexpected additional foliar emissions of CO2 during drought, even after correcting for temperature. Although previously not considered in NEP calculations, this effect has been reported for other trees experiencing low rainfall [Turnbull et al., 2001; Wright et al., 2006] and, as well as reducing NEP, may also increase mortality risk through ex-cessive metabolism of carbon reserves. Other unexpected drought response processes may need to be considered in the future, including the potential for changes in isoprene emissions, currently a small component of the carbon cycle [Pegoraro et al., 2004], in pathogen attacks, known to be substantial during drought stress in other forest ecosystems [Ayers and Lombardero, 2000] and in soil fungal activity [Meir et al., 2006].

6. MODELING TWENTY-FIRST CENTURY DROUGHT IMPACTS ON AMAZONIAN RAIN FORESTS

6.1. Short-Term Effects: Seasonal to Interannual

Notwithstanding the possible discovery of new long-term drought responses, correct attribution of physiological pro-cesses at seasonal or interannual timescales is needed to in-terpret the effects of climate anomalies, such as ENSO or the 2005 drought in Amazonia, and to provide the basis for robust model predictions.

The TFE experiments simulated rainfall reductions similar to that of a severe ENSO, such as the 1997/1998 event [Meir et al., 2008], but the results are only beginning to be incor-porated into vegetation modeling frameworks. Fisher et al. [2007] successfully simulated the effects of the Caxiuanã TFE manipulation on GPP, specifying a detailed multilayer soil and canopy physiological model [Williams et al., 1996] using measurements from the TFE experiment of soil hy-draulic properties, leaf biochemical photosynthetic capacity and LAI, and meteorological data (Figure 6). Gas exchange was validated at leaf and canopy scales using independent leaf-scale stomatal conductance [Fisher et al., 2006] and tree-scale sap flux [Fisher et al., 2007] measurements. The analysis demonstrated that GPP and transpiration at Cax-iuanã are not constrained by moisture supply under normal climatic variation [see Saleska et al., this volume], but that more severe moisture limitation imposed strong constraints upon transpiration and GPP. Transpiration declined by 30–40% (300–418 mm a−1) and GPP by 13–14% (4.0–4.3 t C ha−1 a−1) during the first 2 years of experimental drought at Caxiuanã [Fisher et al., 2007]. Changes in LAI and hydrau-lic (rather than biochemical) properties were the principal determining parameters: maximum foliar stomatal conduct-ance declined by more than 50% and belowground hydrau-lic resistance increased more than 10-fold. Combined with changes in heterotrophic and autotrophic respiration [Meir et al., 2008], the impact of drought on NEP is probably finely balanced and closer to zero than the large carbon emissions predicted by earlier model analyses of the effects on NEP of the 1997/1998 ENSO drought [e.g., Tian et al., 1998; Zeng et al., 2005; Peylin et al., 2005].

Coarser-scale models have also been used to simulate Amazonian drought impacts. Potter et al. [2004] used the Carnegic Ames Stanford Approach (CASA) model to quan-tify rain forest ecosystem functioning during drought, par-tially driving simulations with remotely sensed data. In this work, Rs was more moisture sensitive than in earlier model analyses, and in a subsequent development, the spectral sig-nal from the Tapajós canopy to drought was also success-fully incorporated [Asner et al., 2004], offering the future


prospect of detecting drought stress from space. Recently, in a third-generation development of the SiB ecosystem model [Sellers et al., 1986], Baker et al. [2008] incorporated new soil depth and rooting properties observed at Tapajós [Ol-iveira et al., 2005; Lee et al., 2005], together with a seasonal soil moisture response function for Rs observed at a separate site near the Tapajós TFE experiment [Saleska et al., 2003]. The new model formulation successfully replicated mea-surements of the seasonal variation in forest-atmosphere car-bon exchange made using eddy covariance at Tapajós. Only with the combination of several newly observed mechanisms was it possible to simulate seasonality in carbon exchange correctly with this model [Baker et al., 2008]. However, getting water supply mechanistically correct should ideally start with correctly parameterized soil hydraulic properties, as highlighted by Harris et al. [2004] for a central Ama-zonian site near Manaus [cf. Fisher et al., 2008]. Whether the successful reformulation of SiB by Baker et al. [2008] included a representation of site-corrected soil hydraulic pa-rameters is unclear, but this could strongly influence mod-eled PAW and gas exchange at low soil moisture contents, and underlines the importance of ongoing data-model vali-dation efforts.

6.2. Longer-Term Effects

At longer (decadal-to-century) timescales, it has been necessary to derive the physiological impacts on forest car-bon metabolism of twenty-first century increases in drought stress, temperature, and atmospheric CO2 concentration from physiological principles, as well as from measurements made over shorter time periods [Betts, 2004]. In a recent review, Lloyd and Farquhar [2008] argued that the positive effect of increased atmospheric CO2 concentration on photo-synthesis is likely to outweigh any negative impacts of con-current warming, and this will probably balance in favor of a positive impact on NEP. In addition, the tendency in most plants to reduce stomatal conductance at high atmospheric CO2 concentration makes possible reductions in water loss through transpiration without diminishing carbon acquisi-tion, thus further increasing the resistance of vegetation to climatic drying. However, there remains uncertainty in transpiration estimates, as under drought and/or warming, a drier atmosphere will impose a bigger atmospheric demand on evaporation potentially leading instead to higher rates of evapotranspiration [Salazar et al., 2007] irrespective of reductions in stomatal conductance. Plant respiration [e.g.,

Figure 6. The modeled response to normal rainfall and TFE treatment by GPP and canopy (stomatal) conductance to water vapor, at Caxiuanã: (top) incident rainfall, (middle) stomatal conductance at canopy scale, and (bottom) GPP. Open circles denote TFE treatment forest; solid circles denote control forest. From Fisher et al. [2007], reprinted with permis-sion from Wiley-Blackwell.

MEIR ET AL. 443

Meir et al., 2001] now seems likely to acclimate at higher temperatures [Atkin and Tjoelker, 2003; Atkin et al., 2008], and this would confer drought resistance through reduced use of plant carbon reserves. But whether stomatal conduct-ance responses to increased atmospheric CO2 concentration over the long term remain similar to short-term measure-ments is uncertain, and in any case, the carbon economy of trees and ecosystems may be further influenced by changes in LAI and the drought response in leaf respiration [Atkin and Macherel, 2009; Meir et al., 2008]. Of course, over such longer timescales, any resistance to drought based on plant physiology may also be strongly and negatively impacted by pest or pathogen attack [Ayers and Lombardero, 2000; Meir et al., 2006], or by an increase in the frequency of extreme weather events [IPCC Working Group I, 2007].

Incorporating many of these responses into vegetation models operating over the time periods required to simulate vegetation change is still at an early stage [Meir et al., 2006; Ostle et al., 2009]. However, analysis of the drought response mechanisms specified in different dynamic global vegeta-tion models (DGVMs) has identified the need for corrections to some process representations. The mortality risks recently quantified for natural and more severe drought [Nepstad et al., 2007; Meir et al., 2008; Phillips et al., 2009] have not yet been incorporated into current modeling frameworks and require specific model structures or parameterization to do so [Moorcroft, 2006]. Furthermore, differences exist among different DGVMs in the allocation of fixed carbon to above- and belowground ecosystem components, and this has a sig-nificant impact on the response in Rs to climatic warming and hence to changes in NEP [Dufresne et al., 2002]. Re-cent new insight into carbon allocation processes [see Malhi et al., this volume) should inform this issue further. More surprisingly, Galbraith et al. (submitted manuscript, 2009), analyzing three widely used DGVMs [Cox et al., 2000; Levy et al., 2004; Sitch et al., 2003], have shown that the mod-eled “dieback response” in Amazonian vegetation [Cox et al., 2000; Sitch et al., 2008] is more strongly dependent on the specified temperature responses in respiration and pho-tosynthesis than on the direct effects of moisture limitation, despite observational evidence elsewhere of acclimation to temperature in plant respiration [Atkin and Tjoelker, 2003], and the effects of drought summarized here. Thus, the chal-lenge now is to incorporate the range of observed moisture limitation effects correctly into DGVMs and other vegetation modeling frameworks. Getting the balance right between the drought-buffering effects of above- and belowground ecosystem components and representing them at the correct scale will require a two-way interaction between data pro-viders and modelers. This work will improve the modeling of vegetation-atmosphere interactions during drought, but to

understand the overall effects of drought on Amazonian rain forests, further outward links to models of fire risk, and land use change are also needed.

6.3. Drought and Fire

Switches in vegetation cover under climatic drying and warming have been predicted using both dynamic and equi-librium vegetation model analyses [Oyama and Nobre, 2003; Salazar et al., 2007; Sitch et al., 2008], but none of these simulations is likely to be realistic without the inclusion of fire risk. Lags in the development of natural vegetation un-der climatic change outside the fundamental niche of many tree species may occur because trees are large, resistant, and long-lived organisms. However, fire has the potential to switch forest to savanna or grassland, short-circuiting these lags and rapidly accelerating natural rates of climate-driven vegetation change.

The networks of positive feedbacks among climatic warming and drying, deforestation, forest fragmentation, and fire have been described in detail previously [Nepstad et al., 1999, 2001; Soares-Filho, 2006], and the southeast-ern sector of Amazonia seems most vulnerable to forest loss, as high drought risk and high rates of deforestation overlap [Malhi et al., 2008]. However, dynamic integrated models of climate, fire risk, vegetation, and deforestation have not yet been developed very far [Nepstad et al., 2008]. In one such early study, Golding and Betts [2008] demonstrated substan-tially increased fire risk across the region by 2020, rising to a “high” risk of fire across 50% of the region by 2080. This analysis superimposed deforestation scenarios [Soares-Filho et al., 2006; van Vuuren et al., 2007] and a simple fire model [a forest fire danger index (FFDI), parameterized in Australia] [Noble et al., 1980; Hoffmann et al., 2003] on an ensemble climate model analysis using HADCM3 that incorporated the vegetation response in a simplified way through altered GCM parameter sets [Golding and Betts, 2008]. The next steps in this process will be to incorporate the flammability estimates and vegetation responses from the TFE and other observational data into a fully functioning GCM-DGVM-fire vulnerability framework.

The Australia-derived FFDI model used by Betts [2008] does not consider fire vulnerability in the forest understory, a frequent precursor to subsequent full-canopy fire events in Amazonian forests [Nepstad et al., 2001]. Hence, the re-sults from the TFE experiments (Figure 5) probably indi-cate a higher vulnerability to fire than specified by the FFDI model: under scenarios of stronger drying or greater defor-estation, the biophysical component of the risk to forest loss estimated by Betts et al. [2008] may prove conservative. However, future fire risk is also strongly dependent on the


nature of environmental governance in forested regions be-cause of its impact on deforestation and other modes of land use change [Malhi et al., 2009]. In this regard, there may be some room for optimism. In particular, the potential to mitigate the risk to fire and forest degradation through local and regional governance mechanisms is growing rapidly in parts of Amazonia [Nepstad et al., 2006; Soares-Filho et al., 2006]. The added possibility of national and international agreements that may permit and help finance sustainable land use based on payments for forest ecosystem services [Mitchell et al., 2008; Daily et al., 2009] may further re-duce the risk of drought-related forest loss, thus delaying or avoiding some of the more extreme “dieback” scenarios that have been modeled for the region.

7. CONCLUSIONS

Drought in Amazonia cannot be represented as a single parameter, and current modeling trends are moving toward the representation of the suite of changes in climate, vegeta-tion, soil, fire, and land use that map to this term. The LBA observational network has enriched our capacity to specify such modeling frameworks. The manipulative experiments described here form part of this network and comprise an important means by which we can test the mechanistic ba-sis of the relevant ecological processes, as well as validate measured gross ecosystem fluxes, and thus provide confi-dence in model predictions of future scenarios.

Validating process representation in models at multiyear or longer timescales is becoming increasingly important, and the value of long-term experimental data is increasingly recognized [Sitch et al., 2008]. Combining the experimental results summarized here with observational data from across the basin, a picture of the impacts of drought on Amazonian vegetation is beginning to emerge. This understanding needs to be incorporated into the next generation of DGVMs and can currently be summarized as follows. (1) Resistance to seasonal drought in rain forest functioning is likely at most locations, especially climatically marginal sites. (2) Even un-der severe drought, such as that imposed in the TFE experi-ments, a surprising degree of resistance is probable at first, although spatially patchy increases in mortality were observed during the 2005 drought. In the short term, while strong re-sponses in GPP, transpiration, respiration, and LAI can oc-cur during periods of strong moisture limitation of up to 24 months, overall NEP may change only slightly. (3) As severe drought extends to 3–5 years and beyond, mortality increases markedly, and species differences may emerge in terms of loss, survival, reproduction, and regrowth, with substantial negative impacts on NEP and transpiration, and substantial increases in vulnerability to fire. (4) At longer timescales,

our ability to tightly constrain the tempo and mode of any drought-driven tipping point in forest function is limited by data availability, but the increased risk of fire associated with drought means that switches in vegetation type at decadal or greater timescales will rely as much on human activities as on climate.

New results are still emerging from the three soil mois-ture manipulation experiments at Caxiuanã, Tapajós, and Castanhal. The rainfall exclusion at Tapajós has now been halted, and the processes governing recovery from drought are under analysis [e.g., Davidson et al., 2008]; at Cax-iuanã and Castanhal, the experimental treatments continue. The possibility of long-term data sets offered by these ex-periments represents a uniquely powerful way by which we can begin to understand how multiyear and decadal-scale drought will impact the species composition, vulnerabil-ity, and gas exchange properties of Amazonian vegetation. The standard science funding cycle of 3 or 5 years is in-sufficient to fully address ecological questions of this sort, and although this issue has been recognized within LBA and elsewhere [Hobbie et al., 2003], the case for supporting long-term ecological studies is urgent and needs to be made more widely. As the experiments, data, and model analy-ses are extended, some of the key emerging science chal-lenges are likely to include at least some of the following questions:

1. What is the minimum set of rooting and soil depth properties required to model vegetation and soil function adequately during drought [Woodward and Osborne, 2000; Bruno et al., 2006; Metcalfe et al., 2008]?

2. How can plant hydraulic and biochemical sensitivi-ties be represented accurately at large scale, including their role in affecting tree mortality [Fisher et al., 2006, 2007; McDowell et al., 2008]?

3. How can respiration in plants and microbes be repre-sented over short and long timescales under climatic warm-ing and drying? [Trumbore, 2006; Meir et al., 2008]

4. Can mortality risks and alterations to reproductive out-put be modeled to predict change in vegetation properties and species composition under drought [Brando et al., 2008; Meir et al., 2008; Phillips et al., 2009]?

5. What are the sensitivities in the components of NEP and the allocation of net primary production to PAW and temperature [Brando et al., 2008; Galbraith et al., submitted manuscript, 2009]?

6. Can we model recovery from drought, and how long does recovery take [Brando et al., 2008]?

7. How will twenty-first century land use change, fire inci-dence, and drought interact to affect rain forest functioning or a transition from rain forest to different vegetation types [Soares-Filho et al., 2006; Betts et al., 2008]?

MEIR ET AL. 445

Acknowledgments. We would like to thank several funding bod-ies for the initiation and continuation of research funding support at all three sites from 1999 to 2009: LBA, Brazil; NERC, UK; EU 5th Framework Programme; NASA, USA; NSF, USA; CNPq, Bra-zil. We also thank the institutes responsible for maintenance of the reserves at each site and for providing permission and support for this work: MPEG, Belem, Pará; Santarem, Pará; UFRA, Castanhal, Pará. A large number of Brazilian students have been trained at B.S., M.S., and Ph.D. level during the running of the three experiments, and we thank LBA for making this training support available.

REFERENCES

Alencar, A. A. C., L. A. Solorzano, and D. C. Nepstad (2004), Modeling forest understory fires in an eastern Amazonian land-scape, Ecol. Appl., 14, S139–S149.

Alencar, A. A. C., D. Nepstad, and M. D. V Diaz (2006), Forest understory fire in the Brazilian Amazon in ENSO and non-ENSO years: Area burned and committed carbon emissions, Earth Interact., 10(6), EI150, doi:10.1175/EI150.1.

Aragão, L. E. O. C., Y. Malhi, R. M. Roman-Cuesta, S. Saatchi, L. O. Anderson, and Y. E. Shimabukuro (2007), Spatial patterns and fire response of recent Amazonian droughts, Geophys. Res. Lett., 34, L07701, doi:10.1029/2006GL028946.

Asner, G. P., D. Nepstad, G. Cardinot, and D. Ray (2004), Drought stress and carbon uptake in an Amazon forest measured with spaceborne imaging spectroscopy, Proc. Natl. Acad. Sci. U. S. A., 101, 6039–6044.

Atkin, O. k., and D. Macherel (2009), The crucial role of plant mi-tochondria in orchestrating drought tolerance, Ann. Bot., 103(4), 581–597, doi:10.1093/aob/mcn094.

Atkin, O. k., and M. G. Tjoelker (2003), Thermal acclimation and the dynamic response of plant respiration to temperature, Trends Plant Sci., 8, 343–351.

Atkin, O. k., L. j. Atkinson, R. A. Fisher, C. D. Campbell, j. Zaragoza-Castells, j. W. Pitchford, F. I. Woodward, and V. Hurry (2008), Using temperature-dependent changes in leaf scaling relationships to quantitatively account for thermal ac-climation of respiration in a coupled global climate-vegetation model, Global Change Biol., 14, 2709–2726.

Ayers, M. P., and M. j. Lombardero (2000), Assessing the conse-quences of global change for forest disturbances for herbivores and pathogens, Sci. Total Env., 262, 263–286.

Baker, I. T., L. Prihodko, A. S. Denning, M. Goulden, S. Miller, and H. R. da Rocha (2008), Seasonal drought stress in the Ama-zon: Reconciling models and observations, J. Geophys. Res., 113, G00B01, doi:10.1029/2007jG000644.

Baker, T. R., et al. (2004), Variation in wood density determines spatial patterns in Amazonian forest biomass, Global Change Biol., 10, 1–18, doi:10.1111/j.1529-8817.2003.00751.x.

Belk, E. L., D. Markewitz, T. C. Rasmussen, E. j. M. Carvalho, D. C. Nepstad, and E. A. Davidson (2007), Modeling the effects of throughfall reduction on soil water content in a Brazilian Oxisol under a moist tropical forest, Water Resour. Res., 43, W08432, doi:10.1029/2006WR005493.

Betts, R. A. (2004), Global vegetation and climate: Self-beneficial effects, climate forcings and climate feedbacks, J Phys. IV, 121, 37–60.

Bousquet, P., P. Peylin, P. Ciais, C. Le Quéré, P. Friedlingstein, and P. P. Tans (2000), Regional changes in carbon dioxide fluxes of land and oceans since 1980, Science, 290, 1342–1346.

Brando, P., et al. (2006), Effects of partial throughfall exclusion on the phenology of Coussarea racemosa (Rubiaceae) in an east-central Amazon rainforest, Oecologia, 150, 181–189.

Brando, P. M., D. C. Nepstad, E. A. Davidson, S. E. Trumbore, D. Ray, and P. Camargo (2008), Drought effects on litterfall, wood production and belowground carbon cycling in an Amazon forest: Results of a throughfall reduction experiment, Philos. Trans. R. Soc., Ser. B, 363, 1839–1848.

Bruno, R. D., H. R. da Rocha, H. C. De Freitas, M. L. Goulden, and S. D. Miller (2006), Soil moisture dynamics in an eastern Ama-zonian tropical forest, Hydrol. Processes, 20, 2477–2489.

Cardoso, M. F., G. C. Hurtt, B. Moore, C. A. Nobre, and E. Prins (2003), Projecting future fire activity in Amazonia, Global Change Biol., 9, 656–669, doi:l 0.1046/j.1365-2486.2003.00607.x.

Carswell, F. E., et al. (2002), Seasonality in CO2 and H2O flux at an eastern Amazonian rain forest, J. Geophys. Res., 107(D20), 8076, doi:10.1029/2000jD000284.

Chagnon, F. j. F., and R. L. Bras (2005), Contemporary cli-mate change in the Amazon, Geophys. Res. Lett., 32, L13703, doi:10.1029/2005GL022722.

Cochrane, M. A. (2003), Fire science for rainforests, Nature, 421, 913–919.

Cochrane, M. A., and W. F. Laurance (2002), Fire as a large- scale edge effect in Amazonian forests, J. Trop. Ecol., 18, 311–325.

Coelho, R. F. R., D. G. Zarin, I. S. Miranda, and j. M. Tucker (2004), Análise florística e estrutural de uma floresta em dif-erentes estágios sucessionais no município de Castanhal, Pará, Acta Amazonica, 33, 563–582.

Costa, M. H., S. N. M. Yanagi, P. j. O. P. Souza, A. Ribeiro, and E. j. P. Rocha (2007), Climate change in Amazonia caused by soybean cropland expansion, as compared to caused by pasture-land expansion, Geophys. Res. Lett., 34, L07706, doi:10.1029/2007GL029271.

Cox, P. M., R. A. Betts, C. D. jones, S. A. Spall, and I. j. Totterdell (2000), Acceleration of global warming due to car-bon-cycle feedbacks in a coupled climate model, Nature, 408, 184–187.

Cox, P. M., P. P. Harris, C. Huntingford, R. A. Betts, M. Collins, C. D. jones, T. E. jupp, j. A. Marengo, and C. A. Nobre (2008), Increasing risk of Amazonian drought due to decreasing aerosol pollution, Nature, 453, 212–215.

Culf, A. D., G. Fisch, and M. G. Hodnett (1995), The albedo of Amazonian forest and ranchland, J. Clim., 8, 1544–1554.

Daily, G. C., et al. (2009), Ecosystem services in decision making: Time to deliver, Front. Ecol. Environ., 7, 21–28, doi:10.1890/080025.

da Rocha, H. R., et al. (2009), Patterns of water and heat flux across a biome gradient from tropical forest to savanna in Brazil, J. Geophys. Res., 114, G00B12, doi:10.1029/2007jG000640.


da Rocha, H. R., A. O. Manzi, and j. Shuttleworth (2009), Evapotran-spiration, Geophys. Monogr. Ser., doi:10.1029/2008GM000744, this volume.

Davidson, E. A. (1992), Sources of nitric oxide and nitrous oxide following wetting of dry soil, Soil Sci. Soc. Am. J., 56, 95–102.

Davidson, E. A., L. V. Verchot, j. H. Cattanio, I. L. Ackerman, and j. E. M. Carvalho (2000), Effects of soil water content on soil respiration in forests and cattle pastures of eastern Amazonia, Biogeochemistry, 48, 53–69.

Davidson, E. A., F. Y. Ishida, and D. C. Nepstad (2004), Effects of an experimental drought on soil emissions of carbon dioxide, methane, nitrous oxide, and nitric oxide in a moist tropical for-est, Global Change Biol., 10, 718–730.

Davidson, E. A., et al. (2008), Trace gas emissions Tapajos 5 years, Global Change Biol., 14, 2582–2590.

Dufresne, j. L., P. Friedlingstein, M. Berthelot, L. Bopp, P. Ciais, L. Fairhead, H. Le Treut, and P. Monfray (2002). On the mag-nitude of positive feedback between future climate change and the carbon cycle, Geophys. Res. Lett., 29(10), 1405, doi:10.1029/2001GL013777.

Ehleringer, j. R., et al. (2004), The Third International Conference of the Large-Scale Biosphere Atmosphere Program, Avail-able at ftp://lba.cptec.inpe.br/presentations/LBA-III-Conference-july2004-Brasilia/july28,2004/S15/12_7%20Ehleringer%20Stable%20isotope%20analyses.pdf

Engelbrecht, B. M. j., L. S. Comita, R. Condit, T. A. kursar, M. T. Tyree, B. L. Turner, and S. P. Hubbell (2007), Drought sensitivity shapes species distribution patterns in tropical forests, Nature, 447, 80–82.

Fisher, R. A., M. Williams, R. Lobo do Vale, A. C. Lola da Costa, and P. Meir (2006) Evidence from Amazonian forests is con-sistent with isohydric control of leaf water potential, Plant Cell Environ., 29, 151–165.

Fisher, R. A., M. Williams, A. L. Costa, Y. Malhi, R. F. da Costa, S. Almeida, and P. Meir (2007), The response of an Eastern Amazonian rain forest to drought stress: Results and model-ling analyses from a throughfall exclusion experiment, Global Change Biol., 13, 2361–2378.

Fisher, R. A., M. Williams, M. de Lourdes Ruivo, A. L. Costa, and P. Meir (2008), Evaluating climatic and soil water controls on evapotranspiration at two Amazonian rainforest sites, Agric. For. Meteorol., 148, 850–861.

Foley, j. A., A. Botta, M. T. Coe, and M. H. Costa (2002), El Niño–Southern oscillation and the climate, ecosystems and rivers of Amazonia, Global Biogeochem. Cycles, 16(4), 1132, doi:10.1029/2002GB001872.

Fortini, L. B., S. S. Mulkey, D. j. Zarin, S. S. Vasconcelos and C. R. j. D Carvalho (2003), Drought constraints on leaf gas exchange by Miconia ciliata (Melastomataceae) in the understory of an eastern Amazonian regrowth forest stand, Am. J. Bot., 90, 1064–1070.

Forster, P., et al. (2007), Changes in atmospheric constituents and in radiative forcing, Climate Change 2007: The Physical Science Basis, edited by S. Solomon et al, pp. 129–234, Cambridge Univ. Press, Cambridge, U. K.

Franks, P. j., P. L. Drake, and R. H. Froend (2007), Anisohydric but isohydrodynamic: Seasonally constant plant water gradient

explained by a stomatal control mechanism incorporating vari-able plant hydraulic conductance, Plant Cell Environ., 30, 19–30.

Friedlingstein, P., et al. (2006), Climate-carbon cycle feedback analysis: Results from the C4MIP model intercomparison, J. Clim., 19, 3337–3353.

Furley, P. A., j. Proctor, and j. A. Ratter (1992), Nature and Dy-namics of Forest-Savanna Boundaries, CRC Press, London.

Golding, N., and R. Betts (2008), Fire risk in Amazonia due to climate change in the HadCM3 climate model: Potential in-teractions with deforestation, Global Biogeochem. Cycles, 22, GB4007, doi:10.1029/2007GB003166.

Grace, j., j. Lloyd, j. McIntyre, A. C. Miranda, P. Meir, H. S. Miranda, C. Nobre, j. Moncrieff, j. Massheder, and Y. Malhi (1995), Carbon dioxide uptake by an undisturbed tropical rain forest in southwest Amazonia, 1992 to 1993, Science, 270, 778.

Harris, P., C. Huntingford, P. M. Cox, j. H. C. Gash, and Y. Malhi (2004), Effect of soil moisture on canopy conductance of Ama-zonian rainforest, Agric. For. Meteorol., 122, 215–227.

Hobbie, j. E., S. R. Carpenter, N. B. Grimm, j. R. Gosz, and T. R. Seastedt (2003), The US Long Term Ecological Research pro-gram, BioScience, 53, 21–32.

Hoffmann, W. A., W. Schroeder, and R. B. jackson (2003), Re-gional feedbacks among fire, climate, and tropical deforestation, J. Geophys. Res., 108(D23), 4721, doi:10.1029/2003jD003494.

Houghton, R. A., M. Gloor, j. Lloyd, and C. Potter (2009), The regional carbon budget, Geophys. Monogr. Ser., doi:10.1029/2008GM000718, this volume.

Howard, P. j. A. (1979), Respiration of decomposing litter in rela-tion to temperature and moisture, Oikos, 33, 457–465.

Huete, A., k. Didan, W. van Leeuwen, T. Miura, and E. Glenn (2008), MODIS vegetation indices, Land Remote Sensing and Global Environmental Change: NASA’s Earth Observing Sys-tem and the Science of ASTER and MODIS.

Huntingford, C., j. A. Lowe, B. B. B. Booth, C. D. jones, G. R. Harris, and P. Meir (2009), How large is the effect of model uncertainty in the global carbon cycle compared to model un-certainty in thermal properties of the Earth system when pre-dicting temperature increases by year 2100?, Tellus, Ser. B, 61, 355–360.

Hurlbert, S. H. (2004), On misinterpretations of pseudoreplication and related matters: A reply to Oksanen, Oikos, 104, 591.

Hutyra, L. R., j. W. Munger, C. A. Nobre, S. R. Saleska, S. A. Vieira, and S. C. Wofsy (2005), Climatic variability and vegeta-tion vulnerability in Amazonia, Geophys. Res. Lett., 32, L24712, doi:10.1029/2005GL024981.

IPCC Working Group I (2007), Regional Climate Projections. in Climate Change 2007: The Physical Science Basis. Contribution of Working Group I to the Fourth Assessment Report of the Inter-governmental Panel on Climate Change, edited by S. Solomon et al., pp. 847–940, Cambridge Univ. Press, Cambridge, U. K.

jipp, P. H., D. C. Nepstad, D. k. Cassel, and C. Reis de Carvalho (1998), Deep soil moisture storage and transpiration in forests and pas-tures of seasonally-dry Amazonia, Clim. Change, 39, 395–412.

joslin, j. D., M. H. Wolfe, and P. j. Hanson (2000), Effects of al-tered water regimes on forest root systems, New Phytol., 147, 117–129.

MEIR ET AL. 447

Langenfelds, R. L., R. j. Francey, B. C. Pak, L. P. Steele, j. Lloyd, C. M. Trudinger, and C. E. Allison (2002), Interannual growth rate variations of atmospheric CO2 and its δ13C, H2, CH4, and CO between 1992 and 1999 linked to biomass burn-ing, Global Biogeochem. Cycles, 16(3), 1048, doi:10.1029/2001GB001466.

Lee, j. E., R. S. Oliveira, T. E. Dawson, and I. Fung (2005), Root functioning modifies seasonal climate, Proc. Nat. Acad Sci. U. S. A., 102, 17,576–17,581.

Levy, P. E., M. G. R Cannell, and A. D. Friend (2004), Modelling the impact of future changes in climate, CO2 concentration and land use on natural ecosystems and the terrestrial carbon sink, Global Environ. Change, 14, 21–30.

Likens, G. E., F. H. Bormann, N. M. johnson, D. W. Fisher, and R. S. Pierce (1970), Effects of forest cutting and herbicide treat-ment on nutrient budgets in the Hubbard Brook watershed- ecosystem, Ecol. Monogr., 40(1), 23–47.

Lloyd, j., and G. D. Farquhar (2008), Effects of rising tempera-tures and [CO2] on the physiology of tropical forest trees, Philos. Trans. R. Soc. Ser. B, 363, 1811–1817.

Lloyd, j., j. Grace, A. C. Miranda, P. Meir, S. C. Wong, B. S. Mi-randa, I. R. Wright, j. H. C. Gash, and j. McIntyre (1995), A sim-ple calibrated model of Amazon rain forest productivity based on leaf biochemical properties, Plant Cell Environ., 18, 1129–1145.

Malhi, Y., A. D. Nobre, j. Grace, B. kruijt, M. G. P. Pereira, A. Culf, and S. Scott (1998), Carbon dioxide transfer over a Central Ama-zonian rain forest, J. Geophys. Res., 103(D24), 31,593–31,612.

Malhi, Y., j. T. Roberts, R. A. Betts, T. j. killeen, W. Li, and C. A. Nobre (2008), Climate change, deforestation, and the fate of the Amazon, Science, 319, 169.

Malhi, Y., et al. (2009), Exploring the likelihood and mechanism of a climate-change-induced dieback of the Amazon rainforest. Proc. Natl. Acad. Sci. U. S. A., doi:10.1073/pnas.0804619106.

Malhi, Y., S. Saatchi, C. Girardin, and L. E. O. C. Aragão (2009), The production, storage, and flow of carbon in Amazonian for-ests, Geophys. Monogr. Ser., doi: 10.1029/2008GM000779, this volume.

Marengo, j. A., C. A. Nobre, j. Tomasella, M. D. Oyama, G. S. De Oliveira, R. De Oliveira, H. Camargo, L. M. Alves, and I. F. Brown (2008), The drought of Amazonia in 2005, J. Clim., 21, 495–516.

McDowell, N., et al. (2008), Mechanisms of plant survival and mortality during drought: Why do some plants survive while others succumb to drought?, New Phytol., 178, 719–739.

Meir, P., and j. Grace (2005). The response to drought by tropical rain forest ecosystems, in Tropical Forests and Global Climate Change, edited by Y. Malhi and O. Phillips, pp. 71–80, Oxford Univ. Press, Oxford, U. k.

Meir, P., j. Grace, and A. C. Miranda (2001), Leaf respiration in two tropical rain forests: constraints on physiology by phospho-rus, nitrogen and temperature, Funct. Ecol., 15, 378–387.

Meir, P., P. Cox, and j. Grace (2006), The influence of terrestrial ecosystems on climate, Trends Ecol. Evol., 21, 254–260.

Meir, P., D. B. Metcalfe, A. C. L. Costa, and R. A. Fisher (2008), The fate of assimilated carbon during drought: Impacts on respiration in Amazon rainforests, Philos. Trans. R. Soc. Ser. B, 363, 1849–1855.

Metcalfe, D. B., P. Meir, L. E. O. C. Aragão Y. Malhi, A. C. L. da Costa, A. Braga, P. H. L. Gonçalves, j. de Athaydes, S. S. de Almeida, and M. Williams (2007), Factors controlling spa-tio-temporal variation in carbon dioxide efflux from surface lit-ter, roots, and soil organic matter at four rain forest sites in the eastern Amazon, J. Geophys. Res., 112, G04001, doi:10.1029/2007jG000443.

Metcalfe, D. B., P. Meir, L. Aragão, A. C. L. da Costa, A. P. Braga, P. H. L. Gonçalves, j. de Athaydes Silva jr., S. S. de Almeida, L. A. Dawson, and Y. Malhi (2008), The effects of water avail-ability on root growth and morphology in an Amazon rainforest, Plant Soil, 311, 189–199.

Mitchell, A. W., k. Secoy, N. Mardas, M. Trivedi, and R. Howard (2008), Forests now in the fight against climate change, Forest Foresight Report 1.v3, Global Canopy Programme, Oxford.

Moorcroft, P. R. (2006), How close are we to a predictive science of the biosphere?, Trends Ecol. Evol., 21, 400–407.

Nepstad, D. C., C. R. de Carvalho, E. A. Davidson, P. H. jipp, P. A. Lefebvre, G. H. Negreiros, E. D. da Silva, T. A. Stone, S. E. Trumbore, and S. Vieira (1994), The role of deep roots in the hy-drological and carbon cycles of Amazonian forests and pastures, Nature, 372, 666–669.

Nepstad, D. C., et al. (1999), Large-scale impoverishment of Ama-zonian forests by logging and fire, Nature, 398, 505–508.

Nepstad, D. C., G. Carvalho, A. C. Barros, A. Alencar, j. P. Capo-bianco, j. Bishop, P. Moutinho, P. Lefebvre, U. L. Silva, and E. Prins (2001), Road paving, fire regime feedbacks, and the future of Amazon forests, For. Ecol. Manage., 154, 395–407.

Nepstad, D. C., et al. (2002), The effects of partial throughfall ex-clusion on canopy processes, aboveground production, and bio-geochemistry of an Amazon forest, J. Geophys. Res., 107(D20), 8085, doi:10.1029/2001jD000360.

Nepstad, D. C., P. Lefebvre, U. L. Da Silva, j. Tomasella, P. Schles-inger, L. Solorzano, P. Moutinho, D. Ray, and j. G. Benito (2004), Amazon drought and its implications for forest flammability and tree growth: A basin-wide analysis, Global Change Biol., 10, 704–717.

Nepstad, D., et al. (2006), Inhibition of Amazon deforestation and fire by parks and indigenous lands, Conserv. Biol., 20, 65–73.

Nepstad, D. C., I. Tohver, D. Ray, P. Moutinho, and G. Cardinot (2007), Mortality of large trees and lianas following experimen-tal drought in an Amazon forest, Ecology, 88, 2259–2269.

Noble, I. R., A. M. Gill, and G. A. V. Bary (1980), McArthur’s fire danger meters expressed as equations, J. Ecol., 5, 201–203.

Nobre, C. A., P. j. Sellers, and j. Shukla (1991) Amazonian defor-estation and regional climate change, J. Clim., 4, 957–988.

Oliveira, R. S., T. E. Dawson, S. S. O. Burgess, and D. C. Neps-tad (2005), Hydraulic redistribution in three Amazonian trees, Oecologia, 145, 354–363.

Ostle, N. j., et al. (2009), Integrating plant-soil interactions into global carbon cycle models, J Ecol., 97, 851–863.

Oyama, M. D., and C. A. Nobre (2003), A new climate-vegeta-tion equilibrium state for tropical South America, Geophys. Res. Lett., 30(23), 2199, doi:10.1029/2003GL018600.

Pegoraro, E., et al. (2004), Effect of drought on isoprene emission rates from leaves of Quercus virginiana Mill, Atmos. Environ., 38(36), 6149–6156.


Peylin, P., P. Bousquet, C. Le Quéré, S. Sitch, P. Friedlingstein, G. McKinley, N. Gruber, P. Rayner, and P. Ciais (2005), Mul-tiple constraints on regional CO2 flux variations over land and oceans, Global Biogeochem. Cycles, 19, GB1011, doi:10.1029/2003GB002214.

Phillips, O. L., et al. (2009), Drought sensitivity of the Amazon rainforest, Science, 323, 1344–1347.

Potter, C., et al. (2004), Understanding global teleconnections of climate to regional model estimates of Amazon ecosystem car-bon fluxes, Global Change Biol., 10, 693–703.

Prentice, C., and j. Lloyd (1998), C-quest in the Amazon, Nature, 396, 619.

Ray, D., D. Nepstad, and P. Moutinho (2005), Micrometeorologi-cal and canopy controls of fire susceptibility in a forested Ama-zon landscape, Ecol. Appl., 15, 1664–1678.

Rodenbeck, C., S. Houweling, M. Gloor, and M. Heimann (2003), CO2 flux history 1982–2001 inferred from atmospheric data us-ing a global inversion of atmospheric transport, Atmos. Chem. Phys., 3, 1919–1964.

Ruivo, M., S. B. Pereira, B. Quanz, and P. Meir (2003), Characteri-zation and classification of soils of the LBA experimental sites, Caxiuana, PA, Revista de Ciencias Agrarias, 75–85.

Salazar, L. F., C. A. Nobre, and M. D. Oyama (2007), Climate change consequences on the biome distribution in tropical South America, Geophys. Res. Lett., 34, L09708, doi:10.1029/2007GL029695.

Saleska, S. R., S. D. Miller, D. M. Matross, M. L. Goulden, S. C. Wofsy, H. R. da Rocha, P. B de Camargo, P. Crill, B. C. Daube, and H. C. de Freitas (2003), Carbon in Amazon forests: Unex-pected seasonal fluxes and disturbance-induced losses, Science, 302, 1554–1557.

Saleska, S. R., k. Didan, A. R. Huete, and H. R. da Rocha (2007), Amazon forests green-up during 2005 drought, Science, 318, 612.

Saleska, S., H. da Rocha, B. kruijt, and A. Nobre (2009), Eco-system carbon fluxes and Amazon forest metabolism, Geophys. Monogr. Ser., doi:10.1029/2008GM000728, this volume.

Sampaio, G., C. Nobre, M. H. Costa, P. Satyamurty, B. S. Soares-Filho, and M. Cardoso (2007), Regional climate change over eastern Amazonia caused by pasture and soybean cropland expansion, Geophys. Res. Lett., 34, L17709, doi:10.1029/2007GL030612.

Schwendenmann, L., E. Veldkamp, T. Brenes, j. j. O’Brien, and j. Mackensen (2003), Spatial and temporal variation in soil CO2 efflux in an old-growth neotropical rain forest, La Selva, Costa Rica, Biogeochemistry, 64, 111–128.

Schymanski, S. j., M. Sivapalan, M. L. Roderick, j. Beringer, and L. B. Hutley (2008), An optimality-based model of the coupled soil moisture and root dynamics, Hydrol. Earth Syst. Sci., 12, 913–932.

Sellers, P. j., Y. Mintz, Y. C. Sud, and A. Dalcher (1986), A simple biosphere model (SiB) for use within general circulation models, J. Atmos. Sci., 43, 505–531.

Sitch, S., et al. (2003), Evaluation of ecosystem dynamics, plant ge-ography and terrestrial carbon cycling in the LPj dynamic global vegetation model, Global Change Biol., 9, 161–185.

Sitch, S., et al. (2008), Evaluation of the terrestrial carbon cycle, future plant geography and climate-carbon cycle feedbacks us-ing five Dynamic Global Vegetation Models (DGVMs), Global Change Biol., 14, 25.

Soares-Filho, B. S., D. C. Nepstad, L. M. Curran, G. C. Cerqueira, R. A. Garcia., C. A. Ramos, E. Voll, A. McDonald, P. Lefebvre, and P. Schlesinger (2006), Modelling conservation in the Ama-zon basin, Nature, 440, 520–523.

Sotta, E. D., E. Veldkamp, L. Schwendenmann, B. R. Guima-rães, R. K. Paixão, M. L. P. Ruivo, A. C. L Costa, and P. Meir (2007), Effects of an induced drought on soil CO2 efflux and soil CO2 production in an Eastern Amazonian rainforest, Brazil, Global Change Biol., 13(10), 2218–2229, doi:10.1111/j.1365-2486.2007.01416.x.

Tardieu, F., and T. Simonneau (1998), Variability among species of stomatal control under fluctuating soil water status and evapo-rative demand: Modelling isohydric and anisohydric behaviours, J. Exp. Bot., 49, 419–432.

Tian, H., j. M. Melillo, D. W. kicklighter, A. D. McGuire, j. V. k. Helfrich, B. Moore, and C. j. Vorosmarty (1998), Effect of in-terannual climate variability on carbon storage in Amazonian ecosystems, Nature, 396, 664–667.

Timmermann, A. (1999), Detecting the nonstationary response of ENSO to greenhouse warming, J. Atmos. Sci., 56, 2313–2325.

Tomasella, j., and Hodnett, M. G. (1997), Estimating unsaturated hydraulic conductivity of Brazilian soils using soil-water reten-tion data, Soil Sci., 162, 703.

Trumbore, S. U. S. A, (2006), Carbon respired by terrestrial eco-systems—Recent progress and challenges, Global Change Biol. 12, 141–153.

Trumbore, S., and P. B. de Camargo (2009), Soil carbon dynam-ics, Geophys. Monogr. Ser., doi:10.1029/2008GM000741, this volume.

Turnbull, M. H., et al. (2001), Responses of leaf respiration to tem-perature and leaf characteristics in three deciduous tree species vary with site water availability, Tree Phys., 21, 571–578.

Tyree, M. T., and j. S. Sperry (1988), Do woody plants operate near the point of catastrophic xylem dysfunction caused by dy-namic water stress—Answers from a model, Plant Phys., 88, 574–580.

Uhl, C., and j. B. kauffman (1990), Deforestation, fire susceptibil-ity, and potential tree responses to fire in the eastern Amazon, Ecology, 71, 437–449.

van der Werf, G. R., j. T. Randerson, G. j. Collatz, L. Giglio, P. S. kasibhatla, A. F. Arellano, S. C. Olsen, and E. S. kasischke (2004), Continental-scale partitioning of fire emissions during the 1997 to 2001 El Niño/La Niña period, Science, 303, 73–76.

van Vuuren, D. P., P. L. Lucas, and H. Hilderink (2007), Downs-caling drivers of global environmental change: Enabling use of global SRES scenarios at the national and grid levels, Global En-viron. Change, 17(1), 114–130, doi:10.1016/j.gloenvcha.2006. 04.004.

Vasconcelos, S. S., et al. (2004), Moisture and substrate availability constrain soil trace gas fluxes in an eastern Amazonian regrowth forest, Global Biogeochem. Cycles, 18, GB2009, doi:10.1029/2003GB002210.

MEIR ET AL. 449

S. Almeida, Museau Paraense Emilio Goeldi, Av. Perimetral 1901, Terra Firme, CEP 66077-830, Belém, PA, Brazil.

P. M. Brando and D. Zarin, Department of Botany, University of Florida, Gainesville, FL 32611, USA.

G. Cardinot, Instituto de Pesquisa Ambiental da Amazônia, Av. Nazaré 669, 66035-170, Belém, PA, Brazil.

A. C. L. Costa, Centro de Geociências, Universidade Federal do Pará, Belém, CP 1611 66017-970, Brazil.

E. Davidson, Woods Hole Research Center, Falmouth, MA 02540, USA.

R. A. Fisher, Los Alamos National Laboratory, Los Alamos, NM 87545, USA.

P. Meir, School of Geosciences, University of Edinburgh, Edin-burgh EH8 9xP, Uk. ([email protected])

D. Nepstad, Gordon and Betty Moore Foundation, Palo Alto, CA 94304, USA.

E. D. Sotta, EMBRAPA Amapa, Rod. juscelino kubitscheck km 05, 2600, CEP 68903-419, Macapa, AP, Caixa Postal 10 CEP 68906-970, Brazil.

S. Vasconcelos, EMBRAPA-Amazônia Oriental, Belém, PA CEP 66095-100, Brazil.

Vasconcelos, S. S., D. j. Zarin , M. M. Araújo , L. G. T Rangel-Vasconcelos, C. j. R. De Carvalho, C. L. Staudhammer, and F. A. Oliveira (2008), Effects of seasonality, litter removal and dry-season irrigation on litterfall quantity and quality in eastern Amazonian forest regrowth, Brazil, J. Trop. Ecol., 24, 27–38.

Vourlitis, G. L., j. de Souza Nogueira, N. P. Filho, W. Hoeger, F. Raiter, M. S. Biudes, j. C. Arruda, V. B. Capistrano, j. L. Brito de Faria, and F. de Almeida Lobo (2005), The sensitivity of diel CO2 and H2O vapor exchange of a tropical transitional forest to seasonal variation in meteorology and water availability, Earth Interact., 9, 1–23, doi:10.1175/EI124.1.

Werth, D., and R. Avissar (2002), The local and global effects of Amazon deforestation. J. Geophys. Res., 107(D20), 8087, doi:10.1029/2001jD000717.

West, A. G., k. R. Hultine, j. S. Sperry, S. E. Bush, and j. R. Ehler-inger (2008), Transpiration and hydraulic strategies in a pinon-juniper woodland, Ecol. Appl., 18, 911–927.

White, A., M. G. R. Cannell, and A. D. Friend (1999), Climate change impacts on ecosystems and the terrestrial carbon sink: A new assessment, Global Environ. Change, 9, S21–S30.

Williams, M., et al. (1996), Modelling the soil-plant-atmosphere continuum in a Quercus-Acer stand at Harvard forest: The regu-lation of stomatal conductance by light, nitrogen and soil/plant hydraulic properties, Plant Cell Environ., 19, 911–927.

Williams, M., Y. Malhi, A. D. Nobre, E. B. Rastetter, j. Grace, M. G. P. Pereira (1998), Seasonal variation in net carbon exchange and evapotranspiration in a Brazilian rainforest: A modelling analysis, Plant Cell Environ., 21, 953–968.

Woodward, F. I., and C. P. Osborne (2000), The representation of root processes in models addressing the responses of vegetation to global change, New Phytol., 147, 223–232.

Wright, I. j., P. B. Reich, O. k. Atkin, C. H. Lusk, M. G. Tjoelker, and M. Westoby (2006), Irradiance, temperature and rainfall

influence leaf dark respiration in woody plants: Evidence from comparisons across 20 sites, New Phytol., 169, 309–319.

Zeng, N., A. Mariotti, and P. Wetzel (2005), Terrestrial mecha-nisms of interannual CO2 variability, Global Biogeochem. Cy-cles, 19, GB1016, doi:10.1029/2004GB002273.

The Effects of Drought on Amazonian Rain Forests Effects On... · The functioning of Amazonian rain...

Documents

Transcript of The Effects of Drought on Amazonian Rain Forests Effects On... · The functioning of Amazonian rain...