disturbi in foresta ed effetti sullo stock di carbonio

Transcript

UNIVERSITÀ DEGLI STUDI DI PADOVA
DIPARTIMENTO TERRITORIO E SISTEMI AGRO-FORESTALI
Pubblicazione del Corso di Cultura in Ecologia
ATTI DEL 44° CORSO
DISTURBI IN FORESTA ED EFFETTI SULLO STOCK DI CARBONIO:
IL PROBLEMA DELLA NON PERMANENZA
FOREST DISTURBANCES AND EFFECTS ON CARBON STOCK:
THE NON-PERMANENCE ISSUE
San Vito di Cadore 9-12 Giugno 2008
Disturbi in foresta ed effetti sullo stock di carbonio: il problema della non permanenza
Forest disturbances and effects on carbon stock: the non-permanence issue
Atti del 44° corso di Cultura in Ecologia
Direttore del corso in Cultura in Ecologia: Tommaso Anfodillo.
Organizzazione scientifica del 44° corso: Eva Valese, Elena Dalla Valle, Tommaso Anfodillo.
Sede del corso: Centro Studi per l’Ambiente Alpino, via F. Ossi, 41 – 32046 San Vito di Cadore (BL), tel.
04369311, fax 0436890048, e-mail: [email protected].
Volume a cura di: Eva Valese, Tommaso Anfodillo, Elena Dalla Valle
La presente pubblicazione può essere richiesta presso la segreteria del Dip.to Territorio e Sistemi Agroforestali
dell’Università di Padova o acquisita in formato .pdf al seguente indirizzo web:
http://www.tesaf.unipd.it/Sanvito/atti.htm
La riproduzione anche di parte del volume è consentita purché sia citata la fonte.
Indicazioni per le citazioni bibliografiche
Esempio di citazione di un singolo contributo:
Grassi, G. 2008. Reducing emissions from deforestation in developing countries: the new challenge for climate
mitigation. In Anfodillo T., Dalla Valle E., Valese E., (eds.), Disturbi in foresta ed effetti sullo stock di carbonio:
il problema della non permanenzai, Pubblicazionedel Corso di Cultura in Ecologia, Atti del 44° corso, Università
di Padova: 61-68.
Citazione del volume:
Anfodillo T., Dalla Valle E., Valese E. (eds.) 2008, Disturbi in foresta ed effetti sullo stock di carbonio: il
problema della non permanenza, Pubblicazione del Corso di Cultura in Ecologia, Atti del 44° corso, Università
di Padova.
INDICE
PRESENTAZIONE__________________________________________________________________________
FOREST DISTURBANCES AND EFFECTS ON CARBON STOCK: GENERAL OVERVIEW____________1
Elena Dalla Valle
LE POLITICHE INTERNAZIONALI PER LA GESTIONE DELLE RISORSE FORESTALI E IL PROBLEMA
DELLA NON PERMANENZA: IL RUOLO DEI MERCATI ISTITUZIONALI E DEGLI INVESTIMENTI
VOLONTARI ____________________________________________________________________________17
Davide Pettenella
FOREST FIRES AND AIR POLLUTION______________________________________________________33
Ana Isabel Miranda, C. Borrego
MEDITERRANEAN FIRE REGIMES AND IMPACTS ON FOREST PERMANENCE CASES FROM
NORTHERN CALIFORNIA, NE SPAIN AND CANARY ISLANDS________________________________47
Domingo Molina
REMOTE SENSING FOR POST FIRE FOREST MANAGEMENT__________________________________57
Gherardo Chirici
PERMANENCE OF THE CARBON STOCKS IN THE NORTH AMERICAN BOREAL FOREST UNDER
NATURAL AND ANTHROPOGENIC DISTURBANCE REGIMES_________________________________69
Jean- Francois Boucher Simon GABOURY, Réjean GAGNON, Daniel LORD, Claude VILLENEUVE
WINDTHROW RISK MANAGEMENT. RESULTS FROM ROMANIAN FORESTS___________________77
Ionel Popa
REDUCING EMISSIONS FROM DEFORESTATION IN DEVELOPING COUNTRIES: THE NEW
CHALLENGE FOR CLIMATE MITIGATION__________________________________________________89
Giacomom Grassi
DISTURBANCES IN A CENTRAL EUROPEAN MOUNTAIN FOREST IN THE CONTEXT OF CARBON
DYNAMICS: MORE TEMPORAL DISCONTINUITY THAN EXPECTED___________________________97
Georg. Gratzer , Splechtna B.E. and Rudel, B.
ALLEGATO I:
PHYTOPHAGOUS INSECTS IN THE ENERGY FLOW OF AN ARTIFICAL STAND OF PINUS NIGRA
ARNOLD IN NORTHEN ITALY____________________________________________________________113
Andrea Battisti
Elenco dei Relatori del 44° Corso____________________________________________________________135
Atti dei Corsi di Cultura in Ecologia__________________________________________________________136
PRESENTAZIONE
In sede internazionale già dal 1992, con la Convenzione Quadro sui Cambiamenti Climatici (UNFCC), viene
espressa la preoccupazione su possibili cambiamenti climatici e la necessità di prevedere politiche volte a
mitigarne gli effetti. Il settore agricolo - forestale è stato coinvolto in queste strategie di mitigazione per il
possibile ruolo di sink di carbonio che le piante assolvono con la fotosintesi. Tale azione di assorbimento però
deve essere considerata al netto delle emissioni dovute alla non-permanenza dell’immagazzinamento del
carbonio, che può infatti ritornare in atmosfera a causa di incendi, degradazione delle foreste, attacchi patogeni,
disturbi da vento etc. Per includere tale rischio, nell’ambito del Protocollo di Kyoto, è stato imposto il principio
per cui una volta inserite delle aree agricole o forestali (art. 3.3 e 3.4 PK) nei sistemi di contabilità nazionale, i
paesi (dell’Allegato I) hanno l’obbligo di monitorarne non solo gli assorbimenti, ma anche le emissioni.
I disturbi come fuoco, vento e patologie, influenzano lo sviluppo degli ecosistemi forestali provocando degli
effetti sulle dinamiche dei popolamenti forestali e sulle emissioni di carbonio.
Definire il regime delle perturbazioni in termini di frequenza, estensione ed intensità risulta dunque necessario
per monitorarne e verificarne gli effetti nei diversi ecosistemi forestali per poter così soddisfare anche gli
impegni internazionali.
Per conseguire tali obiettivi, il 44° Corso di Cultura in Ecologia verterà dunque sui disturbi in foresta e sugli
effetti in termini di perdita di carbonio. Tale corso propone, dopo un inquadramento generale degli impegni
derivanti dall’adesione al Protocollo di Kyoto, un approfondimento sull’importanza della valutazione dei disturbi
che interessano le foreste sotto un ottica multidisciplinare. Sarà poi proposto un approfondimento riguardante il
problema degli incendi forestali, in particolare nell’area mediterranea con studi svolti in Portogallo, Spagna ed
Italia. Verranno ripresi ed analizzati nel dettaglio gli aspetti legati ad altri disturbi caratterizzanti in particolare le
regioni della Romania (danni da vento) e le foreste boreali del Canada; verranno presentati poi specifici
contributi per quanto concerne i disturbi biotici e le loro conseguenze in termini di perdita di biomassa. Sarà
effettuata una escursione per analizzare lo sviluppo di una area forestale percorsa da incendio e colpita poi da
attacco da insetti.
In aderenza alla tradizione del Corso di Cultura in Ecologia, gli incontri proposti vorrebbero presentare
risposte concrete alle questioni affrontate perchè S. Vito è sempre stato luogo d’incontro tra il mondo della
ricerca e le diverse figure professionali operanti nel settore forestale.
Tommaso Anfodillo
Eva Valese
Elena Dalla Valle
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GENERAL OVERVIEW
Elena DALLA VALLE
Università degli Studi di Padova - Dipartimento Territorio e Sistemi Agro-Forestali
[email protected]
Elena Dalla Valle
GENERAL OVERVIEW
Introduction
The dominant factor in the radiative forcing of climate in the industrial era is the increasing concentration of
greenhouse gases (GHG) in the atmosphere, of which several occur naturally, but increases in the atmosphere in
the last 250 years are due principally to human activities (Solomon et al. 2007). The concentration of
atmospheric CO2 has augmented from a pre-industrial value of about 280 ppm to 379 ppm in 2005; for the
decade 1995 to 2005 the growth rate of CO2 led to a 20% increase in its radiative forcing; the dominant factors of
increase in atmospheric CO2 are emissions from fossil fuel and the effects of land use change on plant and soil
carbon, and in recent decades emissions have continued to rise (IPCC 2001, Solomon et al. 2007).
Carbon uptake and storage in the terrestrial biosphere arise from the net difference between uptake due to
vegetation growth, changes in reforestation and sequestration, and emissions due to heterotrophic respiration,
harvesting, deforestation, fire, damage by pollution and other disturbance factors affecting biomass and soils.
In many parts of the world the observed changes in biological systems, either anthropogenic or natural, are
coherent over diverse localities and are consistent in direction with the expected effects of regional changes in
temperature; even if there are many difficulties in discerning the effects that are due to adaptation and nonclimatic drivers, the effects of temperature increases have been documented with some confidence on the
alterations to forest disturbances due to fires and pests. Changes have been observed in the type (e.g., fires,
droughts, blowdowns), intensity, and frequency of disturbances that are affected by regional climate change
(either anthropogenic or natural) and land-use practices, and they in turn affect the productivity of and species
composition within an ecosystem, particularly at higher latitudes in the Northern Hemisphere. Frequency of
outbreaks of pests and diseases have also changed, especially in forested systems, and can be linked to changes
in climate (IPCC 2001, 2007).
Despite the intensively managed ecosystem, nature has not been totally mastered and the severe storms that have
happened in these years demonstrate that it is not possible to eliminate the natural disturbance dynamics through
management. On average, 104 million hectares of forest were reported to be significantly affected each year by
forest fire, pests (insects and diseases) or climatic events such as drought, wind, snow, ice and floods. However,
the area of forest affected by disturbances was severely underreported, with information missing from many
countries, especially for forest fires in Africa (FAO/FRA 2006).
In 1990 and 1999 violent storms in Europe damaged 120 and 180 million m3 of wood respectively (UNECE/FAO 2000) and in 1999 seemed to be the first sign of a change in storm patterns, caused by the greenhouse
effect (Schelhaas et al. 2003a).
Schelhaas, analysing data between 1961 and 2000, estimated that damaged wood volume due to forest fire
increased over time from 2.3 million m3 in the period 1961-1970 to 7.4 million m3 in the 1990s. Reported
damage from storms seems to have increased in frequency and magnitude; however the authors underline that
this apparent increase in frequency and intensity is not so unambiguously connected with climate change, but
could also be due to (i) the increase in reported events and (ii) the state of the forest and changes in forest
management over time, given that forest management activities can significantly influence the forest’s resilience
to disturbances. The reported biotic damage also seems to have increased, but less so than the abiotic; also in this
case the authors remark how this increase is affected by the rise in abiotic damage and how the increasing of
growing stock and spread of monocultures augment the pest risk (Schelhaas et al. 2003a). However other authors
(Ayres and Lombardero 2000, Harrington et al. 2001, Bale et al. 2002, Parmesan 2006, Battisti 2008) analysed
the way in which insects react to climate change and assume that an increase in temperature within the vital
limits of a species implies a faster development, even if a lot of different factors have to be analysed (Battisti
2008).
The impact of climate change on ecosystems will differ according to sectors, regions and range of climate
change; the resilience of many ecosystems is likely to be exceeded this century by a combination of climatechange associated disturbances (e.g. flooding, storm, wildfire, insect) and other global change drivers (e.g. landuse change, pollution) (Dale et al. 2001, IPCC 2001, 2007).
Forest disturbances definition
It is hard to define disturbance unequivocally and the definition of what constitute disturbance events varies
among countries; many natural disturbances exercise their effects through reduction in live plant biomass or
rapid change in the pool of actively cycling soil organic matter, but the dividing line between disturbance and
normal function is rather subjective and disturbance is an integral part of the functioning of all ecosystems and
affects most ecosystem processes. Human activities have altered the frequency and magnitude of many natural
disturbances and have also produced new types of disturbance (e.g. large-scale logging).
Chapin et al. (2002) defined disturbance as a relatively discrete event in time and space that alters the structure
of populations, communities and ecosystem and causes changes in resource availability or in the physical
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Forest disturbances and effects on carbon stock: general overview
environment. The same authors also defined disturbance regime as the range of frequency, type, size, timing and
intensity of disturbance characteristic of an ecosystem or region, and disturbance severity as the magnitude of
change in resource supply or environment caused by a disturbance.
Damage to the forest is defined by (UN-ECE/FAO 2000) as a disturbance to the forest which may be caused by
biotic or abiotic agents, resulting in death or a significant loss of vitality, productivity or value of trees and other
components of the forest ecosystem.
For the IPCC Guidelines (IPCC 2003) disturbance is purely a process that reduces or redistributes carbon pools
in terrestrial ecosystems. In 2006 IPCC redefined disturbance as an environmental fluctuation and destructive
event that disturbs forest health, structure, and/or changes resources or physical environment at any given spatial
or temporal scale; disturbances that affect health and vitality include biotic agents such as insects and diseases,
and abiotic agents such as fire, pollution, and extreme weather conditions. They also introduced definitions for
disturbance by diseases (attributable to bacteria, fungi, phytoplasma or virus), by fire (unplanned and
uncontrolled wildland fire) and by insects (insect pests) (FAO/FRA 2006, IPCC 2006).
The impact of disturbance on ecosystem processes depends on its severity, frequency, type, size, timing and
intensity. The severity is the magnitude of change in resource supply or environment caused by disturbance and
it is often influenced by the intensity of the disturbance; the frequency and size are extremely variable among
ecosystems and disturbance types; the timing of disturbance frequently influences its impact and it is often
changed by human activities, in particular for disturbances like grazing, fire, flooding. Disturbance is one of the
key interactive controls that governs ecosystem processes through its effect on other interactive controls; because
of the effect on other controls, changes in disturbance regime alter the structure and functioning of the ecosystem
(Chapin III et al. 2002). Human induced disturbances modify the natural patterns and magnitude of landscape
heterogeneity; even minor variations in topography or edaphic factors can influence the frequency, type or
severity of natural disturbances and disturbance history interacts to influence further disturbance. Any long-term
trend in climate or soil resources that alters disturbance regime will probably alter the characteristic distribution
of patch sizes on the landscape.
Disturbance by forest fire
We can define fire disturbance as caused by wildfire, regardless of whether it broke out inside or outside the
forest/other wooded land. A wildfire is any unplanned and uncontrolled wildland fire that, regardless of ignition
source, may require suppression response (UN-ECE/FAO 2000, FAO/FRA 2006, IPCC 2006).
Fire has been a major factor in the development and management of many of the world’s forests; fires release
most carbon as CO2 and CH4, nitrogen oxides, particulate matter and trace gases are also released (Amiro et al.
2001). Biomass burning contributes up to 50%, 40% and 16% of the total emissions of anthropogenic origin of
carbon monoxide, carbon dioxide and methane respectively (JRC and GVM 2000). Therefore fires not only
impact carbon sequestration by forests, but emit greenhouse gases that potentially affect the climate; this has
some potential positive feedback since GHG driven climate warming may increase fire activity (Flannigan et al.
1998), which increases GHG. Every fire influences the carbon balance for many years and post-fire effects also
cause additional carbon release through decomposition of killed biomass, and different carbon dynamics due to
re-growth of vegetation, vegetation succession, equilibration of soil organic matter and every change to the forest
that affects the strength of the forest carbon sink (Mouillot and Field 2005).
The frequency, size, intensity, seasonality and type of fires depend on weather and climate in addition to forest
structure and composition. Change in climate and fire regimes will have an effect on carbon, nitrogen cycling
and budgets and these could be critical factors in determining if forests are a carbon sink or source on a year-toyear basis (Flannigan et al. 1998, Flannigan et al. 2000, Swetnam and Anderson 2008).
The Global Forest Resource Assessment Report 2005 (FAO/FRA 2006) gathered and organised the available
data on disturbances in all countries; regarding fire there is a relatively greater number of data on intensively
managed semi-natural forests and forest plantations because of the higher investments in monitoring and control
of fire; in natural, fire-dependent forests (e.g. savannah, woodlands and boreal forests), it is more difficult to
assess the true impact of fire. Information is lacking for a number of countries in which forest fires are known to
have occurred. In the period 1998-2002, the average area burned annually was at least 27.7 million ha of forests,
equivalent to 0.9% of the forest area of the reporting countries; an additional 5.1 million ha of other wooded land
were also reported as significantly affected by fire; the highest percentages were reported from Africa and Asia,
while Europe reported the lowest (Fig. 1) (FAO/FRA 2006). The annual average area of forest fires was reported
to have increased in 35 countries, decreased in 31 countries and remained almost constant in 25, so it is difficult
to discriminate any global trends (Fig. 2).
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Elena Dalla Valle
Fig. 1: data from FAO/FRA 2006
Note: As some countries did not report a complete series, figures for 2000 are slightly different from those presented in the preceding table.
Mouillot and Field (2005) reconstructed historical fire maps from 1900 to 2000; they estimated an average 608
million ha per year burned at the end of the 20th century with a gradual increase since the 1950s; forest fires
represent only 11% of the yearly burned area (86% of fires occur in the tropical savannas and grasslands) but
involve more biomass. Forest fires during the early decades of the 20th century were concentrated in the
temperate latitudes, but since the 1970s the centre of forest fire activity has switch to the tropics (Mouillot and
Field 2005).
According to Kondrashov et al. (2004), in the Russian Far East (FE) forests the ecological disturbances
connected with natural resources use by people and industry present a very serious problem; annually the FE
experiences between 2300 and 2500 fires, affecting 300,000-350,000 ha, excluding the years of catastrophic
disaster that occur about every 22 years. Forest fires are acknowledged as one of the most powerful factors
bringing about change in the FE natural environment (Kondrashov 2004). Again in Russia, a study showed that
the fire event densities during seasons of climatic anomalies (2002 and 2003) in the most intact forests were
twice the normal density, but also that the human impact had a constant “multiplication” effect on the fire events
in the non-intact forests (only 13% of the total fire events in non-intact forest is explained by natural disturbance)
4
(Achard et al. 2006, Mollicone et al. 2006). This recent increase in “wild” fires in Eurasian boreal forests has
implications for the global carbon budget (Schiermeier 2005) and so it is important to also take the human role
into account.
Schelhaas et al. (2003) reported an analysis of forest fires in Europe since the 1960s; both the area burned on
forest land only or with other wooded land has increased, specially during the 1980s and early 1990s; the total
burnt area was estimated to be about 0.15% of the total forest area in Europe. In the 1990s the average decreased
slightly, but the variation between years is very wide. Almost half (44%) of the total forest fire area is in Spain
and Portugal and, excluding France, the Mediterranean area accounts for 88%. The average wood volume
damaged annually by forest fires in Europe is 5.5 million m3 for the period 1961-2000. It also increased over
time, like the forest fire area, from 2.3 million m3 in the period 1961-1970 to 7.4 million m3 in the 1990s
(Schelhaas et al. 2003a).
In a study performed in the Klimath-Siskiyou ecoregion (USA), (Staus et al. 2002) reported that between 1972
and 1992 the majority (~ 90%) of this region remained unchanged; cumulative forest disturbance accounted for
10.5% of the forested area of the ecoregion (0.53% annual rate) and only a small portion of mapped disturbance
was due to fire (<1%).
Within the Western Mountain Initiative, Allen et al. (2007) simulated smoke emissions from wildfire across the
western United States; the results suggest increased smoke emissions will affect northern mountains in changing
climates of the future. Based on historical reconstructions and statistical models of 20th century fires, increased
severity and area burned are expected across the Northwest in response to prolonged and more severe droughts
in the 21st century; there will also be a combination of fires, insect outbreaks and possible drought-induced
mortality as is already being seen in the Southwest region (Allen et al. 2007, Swetnam and Anderson 2008).
A report drawn up by Bonnicksen (2008) analyzed four catastrophic California wildfires that together burned
about 57930 ha. The average greenhouse gas emission from combustion caused by the wildfires is about 157 Mg
CO2eq ha-1, and 611 Mg CO2eq ha-1 considering emissions from combustion and decay together (Bonnicksen
2008).
Amiro et al. (2001) reported for Canadian forests, 10771 fires between 1959 and 1999, with an average direct
carbon emission from the large fires of 27 ± 6 Tg C yr-1 (1 Tg = 1012 g), and an individual year range from 3 Tg
C yr-1 in 1978 to 115 Tg C yr-1 in 1995; additional fires of less than 200 ha in size increase these emissions by a
few percent. These direct emissions by forest fires represent a significant part of the carbon budget for Canada
(Amiro et al. 2001). For a large wildfire in Alaska in 2004, Tan et al. (2007) estimated a total C emission of 81.1
± 13.6 Tg burned (3.1 ± 0.7 Kg C m-2), 73% and 27% attributable to consumption of the ground layer and
aboveground biomass respectively; this significant reduction in the ground layer by large fire may result in
heavy impacts on the boreal ecosystem, with an increase in feedbacks between wildfires and climate change
(Tan et al. 2007).
In order to respond to the Kyoto commitment it is important to analyse all aspects of forest fires (size,
seasonality, timing, magnitude) and their changes over time, and also to evaluate biomass and carbon losses and
forest degradation.
Disturbance by insect and diseases
According to the UN-ECE/FAO 2000 definition, disturbances caused by diseases are those attributable to
pathogens such as bacteria, fungi, phytoplasma, or viruses, while disturbances caused by insect pests are
damaging to tree health (IPCC 2006).
Insects and diseases are integral components of forest ecosystems and are normally present at a relatively low
density, but from time to time some species may quickly reach damaging numbers, spatial distribution may
increase and the outbreak may persist for a variable time before subsiding. Climate directly influences the
survival and spread of insects and pathogens, as well as the susceptibility of their forest ecosystems. Changes in
temperature and precipitations affect phytophage and pathogen survival, reproduction, dispersal and distribution
(Dale et al. 2001); indirect consequences of disturbance from phytophages and pathogens include elimination of
nesting trees for birds and negative effects on mycorrhizal fungi (Ayres and Lombardero 2000). Since climate
change can both directly and indirectly influence phytophages and pathogens, the ultimate effects on patterns of
disturbance include increased disturbance in some areas and decreased disturbance in others (Dale et al. 2001).
As stressed in a recent study, changes in climate and pollution inputs will have significant effects on forest
communities, directly altering resource conditions and indirectly affecting disturbances, but these changes are
complex and may vary depending on present community composition and the timing and form of changes in
precipitation patterns (Hurteau and North 2008).
In Europe, Schelhaas et al. (2003) reported that damage from bark beetles is often closely correlated to storm
damage; after a storm, large amounts of damaged wood usually remain in the forest for a long time, providing
suitable microhabitats for the fast growth of the insect population. For the period 1950-2000 the authors
estimated that the average wood volume damaged in Europe was about 2.9 million m3 yr-1 in total. Other biotic
damages were reported mainly from Central European countries (Poland, Germany, Slovak Republic, Czech
Republic and Slovenia).
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Elena Dalla Valle
FAO/FRA 2006 account that globally the combined forest area adversely affected by insects and diseases for the
2000 reporting period (1998-2002) was approximately 68 million ha; the highest area of insect disturbance
reported for a single country was 14.2 million ha in Canada, and 17.4 million ha of disease disturbance in the
USA (Fig. 3, Fig. 4).
Fig. 5 and Fig. 6 show the trends of the data on forest annually affected by diseases and insects for those
countries that provided information for two points in time. The area affected by diseases is slight increasing
globally while the area affected by insects shows a decrease, but no conclusions can be drawn from the data as to
the causative agents or tree species involved and the effects on trees and the forest ecosystem as a whole.
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Fig. 5 : data from FAO/FRA 2006
A lot of scientific literature suggest scenarios by which climate change could significantly alter patterns of
disturbance from forest herbivores, pathogens and insects; probably some types of disturbances will increase,
others will decrease overall and yet others will shift their geographic occurrence (Ayres and Lombardero 2000).
Strong deleterious impacts are possible for forestry economics, biodiversity and landscape aesthetics. There
could be climate feedback from alterations to forest composition and resulting changes in ecosystem attributes
such as carbon pools.
A study carried out to assess bark beetle damages and carbon losses in central Carinthia, Austria (Seidl et al.
2005), stresses how important it is to consider disturbance agents when assessing the carbon sequestration
potential of forest ecosystems. The simulation showed that under climate change conditions bark beetle damage
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Elena Dalla Valle
clearly increases; after the 100-year simulation, standing stock in the 28 sample stands, under climate change
conditions, is about 43% lower than that estimated under current climate due to increased bark beetle damage.
Hurricanes and windstorms
In the FRA Report (FAO/FRA 2006), hurricanes and windstorms are included in the “other disturbances” class,
which includes both abiotic (e.g. wind, snow, ice, floods, tropical storms and drought) and biotic agents (e.g.
camels, beavers, deer and rodents). In Europe the different events are well documented by the EFI’s Database on
Forest Disturbances (Schelhaas et al. 2003b), while for other countries information on disturbances attributed to
these other factors is highly inconsistent.
Ocean temperatures and regional climate events influence the tracks, size, frequency and intensity of hurricanes
(Dale et al. 2001). Tropical cyclones are short-lived phenomena whose impact can range from devastating
floods, high winds and storm surges to beneficial drought-breaking rains and for this reason it is important to
understand frequency, intensity and regions of occurrence in a warmer world (Walsh and Pittock 1998). There is
considerable uncertainty at the regional scale, although evidence supports a modest increase in tropical cyclone
numbers (Walsh and Pittock 1998, USGS 2005, Webster et al. 2005, Chenoweth 2006). Global warming may
accelerate the hydrologic cycle by evaporating more water and producing more intense and possibly more
frequent storms; even if frequency does not increase it is likely that intensity and possibly duration of individual
events will increase because of the warming of the sea and ocean, sources of energy for hurricanes (Walsh and
Pittock 1998, Dale et al. 2001). The effects of hurricanes on the vegetation include rapid and huge tree mortality,
a complex pattern of tree mortality and altered pattern for forest regeneration; these events can also create carbon
sinks by burying woody debris under landslides, transporting wood to environments where wood decomposition
is slower and the rapid biomass accumulation following the event (Lugo 2000).
Different studies have analyzed the possibility of finding a trend in extreme tropical cyclones, trying to
understand if some areas can expect higher event risk for hurricanes over the coming years, whether from natural
cycles, effect of climate change, or both (Nicholls 2004, Meehl et al. 2005, Pielke et al. 2005, Webster et al.
2005, Chan 2006, Landsea et al. 2006, Webster et al. 2006) and not all the authors are in agreement on the
distinguishing of an increased trend in tropical hurricanes, particularly caused by global warming.
Stunfurt et al. (2007) focussed on the effects of hurricanes in the coastal forest to incorporate disturbance into
managed forests. This study stresses the importance of examining forest management objectives with the idea of
incorporating fine-scale disturbance effects such as more complex stand structure into ongoing forest
management in order to reduce vulnerability to damage from future hurricanes. In this way the natural system
could be used to reduce the susceptibility of urban areas and human populations (Stanturf et al. 2007).
Small-scale wind events are products of mesoscale climatic circumstances and may be affected by climate
changes; windstorms can cause heavy mortality, produce canopy disruption, reduce tree density and size
structure, change local environment conditions and cause economic loss to the forest industry (Dale et al. 2001,
Venäläinen et al. 2004). The economic impact of wind damage is mainly severe in managed forests because of
the reduction in the yield of recoverable timber, increased costs and general problems in forest planning. In
addition, broken and uprooted trees left unharvested can lead to detrimental insect attacks on the remaining
growing stock (Peterson 2000, Venäläinen et al. 2004). Thürig et al. (2005) made a 40-year simulation on the
impact of a windthrow on carbon sequestration in Switzerland and the result indicated that forest management
has a strong influence on the carbon budget, while a storm frequency increase of 30% has a small impact on the
national carbon budget of the forest (Thürig et al. 2005).
Peterson et al. (2000) reviewed the existing literature on the occurrence of continental wind-storms and forest
damage in Eastern North America; stand reconstruction from fires cores, fallen trees and tree ring analyses have
revealed that disturbances are not uncommon at the time scale of centuries in NE America, and the return time
for catastrophic wind events varies among locales from a possible low of 1000 years to nearly 2000 years. The
study underlined that the most important tree characteristics influencing damage from wind are size and species
and also that wind damage produces both immediate and longer-term changes in a forest because of the capacity
to alter environmental conditions by increasing light and soil temperature (Peterson 2000).
According to Schelhaas et al. (2003a), the reported damage from storms in Europe seems to have increased since
1850 in both frequency and magnitude, but also in this case it is possible that this increase is due to the
increasing number of reported events. The estimated average annual storm damage over the period 1950-2000 is
18.7 million m3 of wood; most of the damage from storms was reported in the mountainous areas of the SubAtlantic zone, Alpine zone and Central Pannonic zone, with less in the Atlantic region; the Northern zone had a
much lower frequency of accounts of storm damage than Central Europe (Schelhaas et al. 2003a).
A specific study on Swedish forests (Nilsson et al. 2004) analysed recorded storm damage in the period 19012000; 77 individual storms were recorded with an estimated total volume of storm damaged forest of about 110.7
M m3 (cubic meter over bark), distributed over 40 years. Storm damage in Swedish forests has increased during
the last century, with a peak in the 1980s and two major events in 1954 and 1969 that caused 49% of the total
8
damage. The authors also stress that besides the possible changes in wind climate the increased damage could be
explained by an increase in susceptibility brought about by forest management efficiency, with clear-felling,
artificial regeneration and thinning and changes in root architecture, as well as an increased productive forest
area (Nilsson et al. 2004).
In the Global Forest Resource Assessment 2005 (FAO/FRA 2006) for the 2000 period (1998-2002) the annual
average area affected by other disturbances was 8.4 million ha (Fig. 7); the highest area reported for a single
country was 9.3 million ha in Finland; the data reflect a variety of types of disturbances, from catastrophic events
such as hurricanes and wind storms, to longer term, chronic pressures, such as feeding by animals.
The area of other disturbances almost doubled between the two reporting periods in Europe, mainly due to the
effects of severe storms such as those in 1999; wind being a significant factor in Europe and the tropics and
islands for the 2000 reporting period (Fig. 8).
9
Elena Dalla Valle
The IPCC Guidance for the Kyoto Protocol
The carbon cycle includes changes in carbon stocks due to both continuous processes (i.e., growth, decay) and
discrete events (i.e., disturbances like harvest, fire, insect outbreaks, land-use change and other events).
Continuous processes affect carbon stocks in all areas in each year, while discrete events cause emissions and
redistribute ecosystem carbon in specific areas in the year of the event. Disturbances may also have long-lasting
effects, such as decay of wind-blown or burnt trees. The Guidelines (GL) for National Greenhouse Gas
Inventories of the IPCC (IPCC 2006), retrieving the “Good Practice Guidance (GPG) for Land Use, Land-Use
Change and Forestry” (IPCC 2003) gives countries the possibility to choose between three levels (Tier 1, 2 and
3);
Tier 1 methods assume that all post-disturbance emissions (less removal of harvested wood products) are
estimated as part of the disturbance event, and all post-disturbance emissions are estimated in the year of the
event. Under Tier 1, it is also assumed that the average transfer rate into dead organic matter (dead wood and
litter) is equal to the average transfer rate out of dead organic matter, so that the net stock change in this pool is
zero.
There are two fundamentally different and equally valid approaches to estimating stock changes: (i) the processbased approach, which estimates the net balance of additions to and removals from a carbon stock; and (ii) the
stock-based approach, which estimates the difference in carbon stocks at two points in time.
Annual carbon stock changes in any pool can be estimated using the Gain-Loss Method, which can be applied to
all carbon gains or losses. Gains can be attributed to growth (increase of biomass) and to transfer of carbon from
another pool (e.g., transfer of carbon from the live biomass carbon pool to the dead organic matter pool due to
harvest or natural disturbances). Losses can be attributed to transfers of carbon from one pool to another (e.g.,
the carbon in the slash during a harvesting operation is a loss from the aboveground biomass pool), or emissions
due to decay, harvest, burning, etc.
-
Stock-Difference Method (eq.2.5, p. 2.10 GL):
∆C =
(Ct2 − Ct1 )
t 2 − t1
where,
∆C is the annual carbon stock change in the pool (Mg C yr-1),
Ct2 is the carbon in the pool at time t2 (Mg C),
Ct1 is the carbon in the pool at time t1 (Mg C)
-
Gain-Loss Method (eq. 2.4, p.2.9 GL):
∆C= ∆CG – ∆CL
where,
∆C is the annual carbon stock change in the pool (Mg C yr-1),
∆CB is the annual gain of carbon (Mg C yr-1),
∆CL is the annual loss of carbon (Mg C yr-1).
The carbon gain-loss method is applicable to all tiers but the stock-difference method is more suited to Tiers 2
and 3.
To estimate the annual decrease in biomass carbon stocks due to losses (Gain-Loss Method), the following
relationship is shown in eq. 2.11 (p. 2.16 GL); annual biomass loss is the sum of losses from wood removals,
fuelwood removal and other losses resulting from disturbances such as fire, storms, insects and diseases.
∆C L = Lwood −removals + L fuelwood + Ldisturbances
where,
∆CL is the annual loss of carbon (Mg C yr-1)
Lwood-removals is the annual carbon loss due to wood removals (Mg C yr-1),
Lfuelwood is the annual biomass carbon loss due to fuelwood removals (Mg C yr-1),
Ldisturbances is the annual biomass carbon loss due to disturbances (Mg C yr-1)
the following methodology is a generic approach for assessing carbon release following disturbances (eq.2.14,
p.2.18 GL):
10
Ldisturbance = Adisturbance × Bw × (1 + R) × CF × fd
where,
Ldisturbances is the annual biomass carbon loss due to disturbances (Mg C yr-1)
Adisturbance is the area affected by disturbances (ha yr-1)
Bw is the average aboveground biomass of land areas affected by disturbances (Mg d.m. ha-1)
R is the ratio of belowground biomass to aboveground biomass
CF is the carbon fraction of dry matter (Mg d.m.)-1
fd is the fraction of biomass lost in disturbance
The estimate of other losses of carbon requires data on areas affected by disturbances and on the pre-disturbance
biomass stores in the living vegetation, dead wood, forest floor and mineral soil pools of the area; aboveground
biomass estimates of forest type affected by disturbance are necessary, along with belowground biomass to
aboveground biomass ratio and fraction of biomass lost in disturbance.
Under Tier 1 the average biomass default values are given according to the forest types and management
practices; it is assumed that there are no changes of belowground biomass (R=0) and that all Ldisturbance is emitted
in the year of disturbance (fd=1);
Under Tier 2 biomass change due to disturbances will be taken into account by forest category, type of
disturbance and intensity; average values for biomass are obtained from country specific data.
Under Tier 3 it is possible to adopt models which usually utilize spatially referenced information on the year and
type of disturbance.
Tier methods 2 and 3 assume that some of the carbon from the disturbance is emitted immediately and some is
added to the dead organic matter pools (deadwood, litter) or harvested wood products; country specific data are
used.
For inventory reason, changes in carbon stock in biomass are estimated for (i) land remaining in the same landuse category and (ii) land converted to a new land-use category. The reporting convention is that all emissions
and removals associated with a land use change are reported in the new land use category.
The IPCC Guidance: estimating emission from fire
The IPCC Guidelines (2006) provide a generic approach for estimating emissions from fire. Fire is treated as a
disturbance that affects not only the biomass (in particular, aboveground), but also the dead organic matter (litter
and dead wood).
Emissions (CO2 and non-CO2) need to be reported for all fires (prescribed fires and wildfires) on managed lands.
Emissions from wildfires (and escaped prescribed fires) that occur on unmanaged lands do not need to be
reported, unless those lands are followed by a land-use change (i.e., become managed land). CO2 net emissions
should be reported where the CO2 emissions and removals for the biomass pool are not equivalent in the
inventory year, as often happens for forest land if significant woody biomass is killed (i.e., losses represent
several years of growth and C accumulation), and the net emissions should be reported. Factors that decrease the
amount of fuels available for combustion (e.g., from grazing, decay, removal of biofuels, livestock feed) should
be accounted for.
Despite the large natural spatial and temporal variability of fire (in particular that from wildfires), countries
should estimate and report greenhouse gas emissions from fire on an annual basis.
Because of the limited knowledge of the rates of burn formation under contrasting burning conditions and
subsequent turnover rates, the post-fire residues, comprising unburned and partially burnt components, are not
yet considered in the accounting (IPCC 2006).
Conclusion
Taking the whole world and all the disturbances together, about 104 million ha of forests were reported to be
affected each year by forest fire, pests or climatic events in the period 1998-2002; FAO/FRA 2005 underline that
the area of forest affected by disturbances was severely underreported, with information missing from many
countries especially for forest fires in Africa (FAO/FRA 2006).
According to the IPCC Technical Summary of 2007 Report (Parry et al. 2007) the projected impacts of climate
change on the forest and on human health are high, in all continents, but especially in developing countries; in
Africa the loss of “cloud forest” through fire since 1976 has resulted in a 25% annual reduction in water sources
derived from fog. The frequency and extent of forest fires in northern Asia are expected to increase in the future
due to climate change and extreme weather events that would likely limit forest expansion. In Australia and New
Zealand an increase in forest fire danger is likely, production from agriculture and forestry is projected to decline
over much of southern and eastern Australia and over parts of eastern New Zealand, due to increased drought
and fire; however in New Zealand benefits are projected in western and southern areas and close to major rivers
due to a longer growing season, less frost and increased rainfall. In Europe the forested area is likely to increase
11
Elena Dalla Valle
in the north and decrease in the south; a redistribution of tree species and a rise of the mountain treeline is
expected; nevertheless forest fire risk is almost certain to greatly increase in southern Europe. In Latin America
the frequency and intensity of hurricanes are likely to increase, especially in the Caribbean Basin; increases in
temperature and decrease in soil water would lead to a replacement of tropical forest in Eastern Amazonia and
southern Mexico. In north America climate change in the early decades of the 21st century is likely to increase
forest production, but with high sensitivity to drought, storms, insects and other disturbances; by the second half
of the century the greatest impacts on forests will possibly be through changing disturbances from pests, diseases
and fire, with an extended window of high fire risk and area burned. As stressed by Kurz and Apps (1999),
changes in forest disturbance regimes have resulted in an increase of the average age of Canadian forests over
the period 1920–1980; these changes have also resulted in an increase of the C content of both biomass and
DOM C pools. The primary factor underlying the C sink to ~ 1970 was a change in the disturbance frequency
over a timescale of many decades, which resulted in changes in the average forest age and C content. As the
average forest age increases, the ability to sequester additional C decreases, and the susceptibility to disturbances
increases. Indeed, the changes in the last two decades of the analysis period, whether due to human-induced
climate change or natural variation, have resulted in significant changes in the disturbance regimes, relative to
the preceding half century; these disturbances have decreased the forest sector C sink and resulted in a net
decline in ecosystem C and in a release of C to the atmosphere (Kurz and Apps 1999, Kurz et al. 2008).
In the Polar regions, large-scale forest fires and outbreaks of tree-killing insects that are characteristic of the
boreal forest and some tundra with trees areas are also likely to increase (IPCC 2007, Parry et al. 2007).
A recent study shows us worrying simulations made for Canada’s managed forests; taking into account future
disturbances the authors projected that these forests could be a source of between 30 and 245 million of Mg CO2e
yr-1 during the first Kyoto Protocol commitment period, due to large insect outbreaks and large fires (2008-2012)
(Kurz et al. 2008). Other forested areas, currently estimated to be carbon sinks, may also switch to sources if the
impact of disturbances increases.
With respect to lasting major changes in carbon stocks in forests it is important to focus on stand replacing
disturbance events, which lead to a rapid relocating of carbon from live trees to litter or from live trees and soil
to the atmosphere. A comprehensive analysis should consider in detail the distribution of existing forests and
their structure, as well as the loss of carbon due to disturbances (Goodale et al. 2002, Magnani et al. 2007).
The task of translating areas affected by disturbances into global carbon budgets is difficult to achieve; useful
estimates will include not only the carbon emitted to the atmosphere during the events, but also the carbon
dynamics reflecting degradation, decomposition and re-growth, and for fire also potential combustion in future
fires. Long-term analyses of disturbances and their consequences have to be considered in carbon balance
studies.
Another critical topic regarding forest disturbances assessment and carbon stock conservation is the involvement
of the developing countries in the Kyoto Protocol system, not only in the Clean Development Mechanism
(CDM) projects (that in the land-use sector foresee only afforestation – reforestation projects), but also
supporting forest management and forest conservation, in order to avoid deforestation. In particular, developing
countries could make a substantial contribution to climate change mitigation through native forest conservation
or reduction of deforestation or fires, but they need specific provisions or financial incentives that would commit
these countries to participating in the Kyoto efforts though forest conservation (Moutinho et al. 2005).
12
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15
16
LE POLITICHE INTERNAZIONALI PER LA GESTIONE DELLE RISORSE
FORESTALI E IL PROBLEMA DELLA NON PERMANENZA: IL RUOLO DEI
MERCATI ISTITUZIONALI E DEGLI INVESTIMENTI VOLONTARI
Davide Pettenella
Università degli Studi di Padova - Dipartimento Territorio e Sistemi Agro-Forestali
[email protected]
Davide Pettenella
LE POLITICHE INTERNAZIONALI PER LA GESTIONE DELLE RISORSE FORESTALI E IL
PROBLEMA DELLA NON PERMANENZA: IL RUOLO DEI MERCATI ISTITUZIONALI E DEGLI
INVESTIMENTI VOLONTARI
Introduzione
Il tema della permanenza delle foreste può essere trattato sotto diversi punti di vista; in questo lavoro viene
privilegiato quello relativo alle politiche internazionali volte a consentire direttamente la permanenza degli
ecosistemi forestali. Le cause indirette, essenzialmente connesse alla povertà e agli interessi economici legati
allo sfruttamento irrazionale delle foreste, fanno da cornice all’analisi effettuata, senza tuttavia essere, per ovvie
ragioni di spazio, oggetto di approfondimento.
Il testo è organizzato in tre parti: nella prima viene presentata, in termini molto sintetici, un’analisi dello stato
delle risorse forestali a livello internazionale, con un approfondimento dei problemi legati alla deforestazione,
propedeutico all’analisi dei problemi della non permanenza. Nel secondo capitolo l’attenzione è posta sulle
politiche promosse dalle istituzioni internazionali per la tutela delle risorse forestali. Infine, nel terzo capitolo,
vengono presi in considerazione gli impatti specifici delle politiche di tutela del clima sulla permanenza delle
risorse forestali, approfondendo il ruolo degli investimenti di tipo volontario.
1. La dinamica delle risorse forestali
Le foreste coprono circa il 30% della superficie terrestre e sono una componente essenziale per il mantenimento
della stabilità ambientale: le foreste sono gli ecosistemi più ricchi in termini di biodiversità, hanno un ruolo
fondamentale nel ciclo dell’acqua, nella prevenzione dell’erosione e delle frane, nella fissazione di anidride
carbonica, il principale gas responsabile dell’ “effetto serra” e quindi nella prevenzione dei fenomeni di
riscaldamento globale.
Le foreste hanno un ruolo chiave anche nell’economia di molti paesi: 1,2 miliardi di persone vivono sfruttando le
foreste per coprire fabbisogni essenziali (Mery et al., 2005); 240 milioni vivono in ambienti forestali o
nell’immediata prossimità (WB, 2002); 60 milioni di indigeni hanno nelle foreste la loro quasi esclusiva fonte di
vita. Il legname è la fonte energetica di base per le esigenze di cottura dei cibi e il riscaldamento per 2 miliardi di
persone e copre più del 70% del fabbisogno energetico delle popolazioni africane e del sud-est asiatico. Il ruolo
delle attività economiche connesse alla estrazione e lavorazione del legname non è particolarmente significativo
e il contributo del settore nell’economia mondiale si sta riducendo (dall’1,6 all’1,2% del Prodotto Interno Lordo
mondiale (FAO, 2005), 47 milioni di occupati diretti), ma le foreste mantengono un ruolo importante per le
attività ad esse collegate, spesso basate sull’economia informale: la caccia e la raccolta di frutti, piante
selvatiche, erbe medicinali e aromatiche; l’estrazione di lattici, resine, corteccia e sughero; il turismo e
l’educazione ambientale, ecc. In 62 paesi del Terzo Mondo la caccia in foresta copre più del 20% del fabbisogno
proteico della popolazione rurale (Bennett e Robinson, 2000); in Asia 1,8 miliardi di persone utilizzano piante
selvatiche a scopo medicinale (Srivastava et al., 1996).
Nonostante ci sia un ampio riconoscimento delle funzioni positive associate alla permanenza delle foreste, i
processi di deforestazione non si sono andati riducendo in questi ultimi anni: in base ai dati della FAO1 , nel
periodo 1990-2000 sono stati persi annualmente 14,6 milioni di ettari di foreste naturali (0,38% della superficie
mondiale che è pari a 3,8 miliardi di ettari) e 1,5 milioni di ettari sono stati convertiti in piantagioni, una perdita
solo in parte compensata da 3,6 milioni di ettari di espansione naturale del bosco su terreni abbandonati, e da 3,1
milioni di ettari di nuove piantagioni forestali (FAO, 2000). I dati più recenti, sempre di fonte FAO, confermano
questi trend: secondo l’inventario forestale mondiale del 20052, nell’ultimo quinquennio la diminuzione media
annua della superficie forestale è stata di 13 milioni di ettari.
Il quadro problematico è, tuttavia, peggiore di quanto emerga dall’analisi dei dati statistici di fonte ufficiale: la
deforestazione si basa infatti su una definizione di foresta adottata dalla FAO molto ampia (basta una copertura
delle chiome del 10% di un terreno perché questo sia classificato come foresta – vd. allegato). Sia per il fatto che
ormai in molti paesi le foreste naturali si sono ridotte all’osso, sia perché fenomeni di deforestazione radicale e
su ampie superfici tendono ad essere più controllati, il problema si identifica soprattutto con i processi di
progressivo degrado delle foreste, un processo più difficile da monitorare e controllare. In base al recente
rapporto pubblicato dall’International Tropical Timber Organisation3, nel 2005 solo il 7% dei 352 milioni di
ettari di foreste primarie dei tropici sono gestite per la produzione sostenibile del legname.
Le cause del degrado delle foreste sono state indagate da diversi autori e istituzioni (Buttoud, 2001; Mery et al.,
2005; Scotland e Ludwig, 2002). Ovviamente è impossibile una generalizzazione, anche per la presenza di fattori
molteplici che agiscono in stretta successione o parallelamente. Una foresta statale data in concessione
ventennale ad una impresa privata per la produzione di legname a fini commerciali può essere oggetto di un
1
Cfr. < http://www.fao.org/forestry/fo/fra/main/index.jsp >
Cfr. < http://www.fao.org/docrep/008/a0400e/a0400e00.htm >
3
Cfr. < http://www.itto.or.jp/live/PageDisplayHandler?pageId=270 >
2
18
Le politiche internazionali per la gestione delle risorse forestali e il problema della non permanenza: il ruolo dei
mercati istituzionali e degli investimenti volontari
intervento selettivo di prelievo da parte del concessionario, ma le infrastrutture da questo realizzate possono
essere utilizzate per il successivo prelievo informale di altro legname da parte dei locali o di piccole imprese
irregolari. La presenza di squadre di taglio ed esbosco in foreste primarie comporta spesso il bracconaggio, una
modalità molto diffusa di integrazione dei redditi dei lavoratori forestali.
La Commissione Europea ricorda che «lo sfruttamento illegale delle risorse naturali, incluse quelle forestali, è
strettamente legato alla corruzione e alla criminalità organizzata. In alcuni paesi provvisti di vaste risorse
forestali la corruzione, alimentata dai proventi dei disboscamenti illegali, è arrivata ad un punto tale da
rappresentare una minaccia per lo Stato di diritto, i principi di una governance democratica e il rispetto dei diritti
umani» (CE, 2003). Il taglio illegale e l’uso dei profitti da questi derivanti è, in altri termini, uno dei più evidenti
esempi di degenerazione dei processi di gestione della cosa pubblica assommando e integrando in sé problemi
connessi a corruzione, conflitti sociali, inefficiente uso delle risorse, distrazione di risorse pubbliche a fini
privati.
La riduzione della superficie forestale è spesso conseguenza dell’espansione dell’agricoltura commerciale e di
sussistenza: la coltivazione intensiva con tecniche che portano al rapido degrado della fertilità del suolo, il
sovrapascolamento, l’impiego delle pratiche tradizionali del taglio, incendio e coltivazione temporanea delle aree
forestali (slash and burn, cioè «taglia e brucia») secondo sistemi di rotazione nell’uso dei terreni forestali
divenuti insostenibili a causa dei brevi periodi di riposo e ricostruzione del manto forestale.
Deforestazione e degrado delle foreste non sempre peraltro si identificano con violazione delle norme locali di
settore. Sist et al. (2003) hanno dimostrato, ad esempio, che la norme statali definite in molti paesi del sud-est
asiatico relative ai limiti minimi dei diametri delle piante da tagliare non comportano una gestione sostenibile
delle foreste. Non di rado foreste primarie sono state distrutte sulla base di programmi governativi di
colonizzazione agricola, di trasferimento di agricoltori senza terra, anche per il controllo politico di un territorio,
o sono state trasformate in piantagioni industriali per la coltivazione di specie come eucalipti, pini, l’albero della
gomma, il teak, ecc.
Secondo Transparency International (2004), un’autorevole organizzazione non governativa impegnata nella lotta
a lungo termine contro la corruzione, il grado di corruzione nel settore forestale è nella media rispetto ad altri
settori economici. È invece inferiore rispetto a settori critici quali gli appalti pubblici, la compravendita di
armamenti, gli investimenti in campo energetico. Una grande differenza tra settore forestale e altri settori si
gioca sulle conseguenze dell’illegalità sulle risorse naturali: in una sorta di circolo vizioso, l’eccessivo
sfruttamento determina l’esaurimento delle risorse naturali dalle quali le popolazioni dei paesi in via di sviluppo
dipendono, una dipendenza che viene potenziata dalle condizioni di povertà e rapida crescita demografica. La
correlazione tra deforestazione e pressione demografica è stata evidenziata da vari studi (Repetto e Holmes,
1983; Palo e Salmi, 1987; Palo 1990): l’incremento della popolazione incide sui consumi energetici con
l’aumento dei prelievi di legna da ardere, sulla domanda di aree da destinare alle coltivazioni e all’allevamento
del bestiame (transumanza e agricoltura itinerante che diventano insostenibili; agricoltura da esportazione
estensiva con effetti di rapido depauperamento dei terreni), sulla domanda di aree abitative e da destinare alla
costruzione di strade e infrastrutture e, quindi, sui consumi di legname da costruzione. Le conseguenze negative
dell’accresciuta pressione demografica sulle foreste sono talvolta molto evidenti, come ad esempio si è avuta
testimonianza con il cataclisma dello tsunami del 2004 nell’Oceano indiano: dove non erano state distrutte o
impoverite, le foreste di mangrovia sono state in molti casi, secondo le interpretazioni più autorevoli, una
barriera efficace contro questo evento climatico estremo. Gli effetti dei prelievi industriali di legname possono
anche molto velocemente portare elementi di squilibrio nella popolazione locale: la presenza occasionale di
squadre di operai forestali è, ad esempio, la causa principale di diffusione dell’AIDS nelle comunità rurali più
isolate del Camerun.
Da ultimo, è importante ricordare che le ricche foreste tropicali rappresentano un capitale facilmente
mobilizzabile per esigenze finanziarie estranee all’economia rurale: il taglio e la vendita illegali del legname, ad
esempio, sono fonte di finanziamento di conflitti armati e di commercio di armi. Il problema è così grave che il
Consiglio di Sicurezza dell'ONU ha coniato un termine specifico per caratterizzarlo: il “legname da guerra”
(conflict timber), con il quale ci si riferisce al commercio di prodotti legnosi gestito da gruppi armati, da fazioni
ribelli, da militari o dalle autorità governative allo scopo specifico di alimentare un conflitto o di trarre vantaggi
e profitti illeciti dal conflitto stesso. La Repubblica Democratica del Congo, la Cambogia, la Birmania, la Sierra
Leone e la Liberia sono tra i paesi dove c’è stata maggior evidenza del problema (Global Witness, 2002).
Il commercio di legname è, quindi, solo una delle cause di deforestazione, e non sempre la principale, ma
certamente quella collegata a una maggior responsabilità diretta tra i consumatori e le imprese occidentali.
2. Le politiche internazionali volte alla tutela delle foreste
A fronte di questi problemi, a partire dagli anni ’70 le iniziative della società civile organizzata a favore della
tutela delle foreste si sono andate moltiplicando e perfezionando passando dalla fase di denuncia a quella dei
boicottaggi, fino ad arrivare alla proposta e messa in pratica di strumenti più avanzati ed efficaci: la creazione di
sistemi di monitoraggio, la definizione di criteri e indicatori di buona gestione forestale, la certificazione
dell’origine del legname e della sua tracciabilità, le politiche di acquisto responsabile da parte della pubblica
amministrazione, ecc.
19
Davide Pettenella
Con un certo ritardo rispetto a queste iniziative, anche le istituzioni internazionali hanno avviato una serie di
programmi basati su accordi vincolanti per i paesi, ma soprattutto su soft law, cioè su impegni generici e non
impegnativi sul piano giuridico. Le ragioni di questo ritardo sono diverse: oltre alla normale inerzia nei processi
di negoziazione internazionale, l’assunzione di impegni per fermare i processi di deforestazione è stata rallentata
da tre fattori:
• lo stretto rapporto esistente tra povertà e degrado delle foreste: come impedire l’accesso al legname,
l’unica fonte energetica per la stragrande maggioranza della popolazione che vive nelle aree rurali dei
paesi in via di sviluppo?
• il rispetto del principio dell’autodeterminazione nella definizione delle linee di sviluppo nazionale:
perché i paesi tropicali non dovrebbero fare quello che i paesi europei hanno fatto nella loro storia meno
recente, ovvero distruggere il proprio patrimonio forestale per promuovere lo sviluppo agricolo e
industriale?
• il ruolo fondamentale che, nei processi di deforestazione, ha la piccola e grande corruzione, un tema
sensibile per le autorità di molti governi, a partire da quelli in via di sviluppo4.
L’insieme delle norme internazionali volte alla tutelare la permanenza delle risorse forestali si è andato
arricchendo significativamente a partire dal 1992, l’anno della conferenza delle Nazioni Unite su Ambiente e
Sviluppo (UNCED), nota anche come Earth Summit di Rio. A conclusione della Conferenza, sono stati approvati
una serie significativa di documenti relativi alla gestione delle risorse forestali (Humphreys, 1996),
sinteticamente richiamati nel Quadro 1.
Sul piano della ristrutturazione del quadro istituzionale, conseguenza molto significativa di Rio è stata la
creazione, presso la Commissione per lo Sviluppo Sostenibile delle Nazioni Unite, dell'Intergovernmental Panel
for Forests (IPF), organismo con un mandato a termine (1995-97) che è stato sostituito ora
dall'Intergovernmental Forum for Forests (IFF) con il compito di rappresentare un tavolo di trattativa e di
coordinamento delle politiche di gestione delle foreste (cui partecipano anche organismi non governativi). Nella
definizione del proprio piano di lavoro, l’IFF si avvale dell’Interagency Task Force on Forests (ITFF), un
organismo informale di coordinamento tra tutte le istituzioni delle Nazioni Unite che si occupano di risorse
forestali, e di un Forest Adviser Group. L’IFF ha tenuto la sua ultima sessione alla fine del gennaio 2000,
arrivando ad approvare una serie di raccomandazioni, tra le quali la definizione di un International Arrangement
on Forests e la creazione dell’United Nations Forum on Forests (UNFF), ovvero di un quadro coerente di
normative internazionali di settore e di un organismo di raccordo inter-istituzionale che funga anche da tavolo di
negoziazione di iniziative di sviluppo regionale e locale.
Quadro 1 - I documenti relativi al settore forestale adottati nella Conferenza di Rio su Ambiente e
Sviluppo.
I Principi Forestali5: norme generali di buon comportamento che coprono ogni aspetto della gestione forestale.
Viene affermata la responsabilità di ogni paese nella gestione delle proprie foreste, il ruolo del settore forestale
nello sviluppo economico, la necessità della difesa della biodiversità e la responsabilità – anche in termini
finanziari – di tutti i paesi nel conservare le risorse boschive.
Agenda 21 – Capitolo 11 “Contrastare il disboscamento”: sono evidenziate una serie dettagliate di misure
volte a contrastare i processi di disboscamento, cercando di colpire le cause prime del fenomeno collegate alle
condizioni di povertà, agli interessi commerciali di breve periodo, alla debolezza delle istituzioni, alla mancato
coinvolgimento di tutti i soggetti responsabili della gestione forestale.
In altri capitoli di Agenda 21 si fa esplicito riferimento ad interventi connessi alla gestione sostenibile delle
foreste: il Capitolo 12 “Contrastare la desertificazione e la siccità”, il Capitolo 13: “Lo sviluppo sostenibile delle
aree di montagna”, il Capitolo 15 “La conservazione della diversità biologica”. Le attività forestali, come
opzione alternativa nelle politiche di uso del territorio, sono menzionate nel Capitolo 10 “Un approccio integrato
alla pianificazione e gestione delle risorse territoriali” e nel Capitolo 14 “Uno sviluppo rurale e dell’agricoltura
sostenibili”.
La Convenzione Quadro sui Cambiamenti Climatici: considera gli interventi di monitoraggio e prevenzione
delle emissioni di gas serra, riconoscendo anche al settore forestale un ruolo significativo nelle strategie volte a
ridurre i cambiamenti climatici.
La Convenzione sulla Biodiversità: viene riconosciuto il grande valore per la collettività e la responsabilità per
i paesi nel mantenimento della diversità biologica in tutte le sue diverse forme. Vengono ipotizzate diverse
attività, da attuare tramite piani nazionali, volte a conservare e migliorare la biodiversità: gestione sostenibile
4
Solo nel 2001, con la pubblicazione del secondo rapporto della FAO sullo stato delle risorse forestali mondiali, è stata
utilizzata in un documento ufficiale la locuzione “tagli illegali” (illegal logging) in sostituzione di quelle impiegate in
precedenza che facevano riferimento a concetti come “gestione non sostenibile” o “foreste gestite senza rispetto della
legislazione nazionale”.
5
La denominazione completa dei Principi è: Non-legally Binding Authoritative Statement of Principles for a Global
Consensus on the Management, Conservation and Sustainable Development of All Types of Forests.
20
delle risorse naturali, ricerca, formazione, educazione, accesso pubblico ai benefici derivanti dal mantenimento
del patrimonio genetico, trasferimento di tecnologia e di fondi di investimento, ecc.
La Convenzione per la Lotta alla Desertificazione6: vengono previste una serie di misure, da attuare tramite
piani nazionali, per contrastare i fenomeni di inaridimento. Tra le misure ipotizzate gli interventi che riguardano
le risorse forestali hanno un particolare rilievo.
Ma altre significative iniziative erano già state avviate prima dell'Earth Summit di Rio del 1992. In Europa un
peso significativo nell’integrazione e coordinamento delle politiche di settore viene svolto tramite le Risoluzioni
delle Conferenze Interministeriali per la Protezione delle Foreste in Europa. Le conferenze hanno portato
all’approvazione di una serie di Risoluzioni che impegnano direttamente i Ministeri dei paesi europei con
responsabilità nel settore forestale (vd. Quadro 2).
Sempre prima della Conferenza di Rio, su stimolo della FAO e di altri organismi internazionali erano stati
avviati i Tropical Forest Action Plan per l'impostazione di politiche forestali sostenibili nei paesi in via di
sviluppo e il coordinamento degli interventi di cooperazione multi- e bi-laterale. L’esperienza dei Tropical
Forest Action Plan non è stata esente da critiche, soprattutto da parte delle organizzazioni non governative, ma
ha portato ad una forte attenzione non solo rispetto alla necessità che ogni paese (compresi quelli ad alto tasso di
sviluppo) si doti di strumenti di programmazione settoriale, ma anche alle modalità procedurali per la definizione
dei piani forestali nazionali che devono essere basati sulla più ampia partecipazione e concertazione delle parti
interessate. Su queste tematiche, la FAO, l’IPF prima e l’IFF successivamente, nonché alcune istituzioni non
governative si sono particolarmente impegnate in un’opera di sensibilizzazione e di concreto sostegno delle
esperienze nel settore.
Quadro 2 – Le risoluzioni delle Conferenze Interministeriali per la Protezione delle Foreste in Europa.
Nelle Conferenze sono state approvate, oltre a diverse Dichiarazioni generali, le seguenti Risoluzioni (la prima
lettera della sigla fa riferimento alla sede della conferenza):
Conferenza di Strasburgo (1990):
• Risoluzione S1: Rete europea di punti campione permanenti per il monitoraggio degli ecosistemi forestali;
• Risoluzione S2: Conservazione delle risorse genetiche forestali;
• Risoluzione S3: Banca dati europea a livello decentrato relativa agli incendi boschivi;
• Risoluzione S4: Adattamento delle forme di gestione delle foreste di montagna alle nuove condizioni
ambientali;
• Risoluzione S5: Sviluppo della rete di ricerca EUROSILVA sulla fisiologia degli alberi;
• Risoluzione S6: Rete europea per la ricerca negli ecosistemi forestali.
•
•
•
•
Conferenza di Helsinki (1993):
Risoluzione H1: Linee guida generali per la gestione forestale sostenibile in Europa;
Risoluzione H2: Linee guida generali per la conservazione della biodiversità nelle foreste europee;
Risoluzione H3: La cooperazione in campo forestale con in paesi con economie in transizione;
Risoluzione H4: Strategie per un processo di adattamento delle foreste europee ai cambiamenti climatici.
Conferenza di Lisbona (1998):
• Risoluzione L1: La popolazione, le foreste e il settore forestale. Il miglioramento degli aspetti socioeconomici della gestione forestale sostenibile;
• Risoluzione L2: Criteri, Indicatori e Linee guida a livello operativo per la gestione forestale sostenibile a
livello pan-europeo.
•
•
•
•
•
Conferenza di Vienna (2003):
Risoluzione V1: Cooperazione intersettoriale e programmi forestali nazionali
Risoluzione V2: Fattibilità economica della gestione forestale sostenibile
Risoluzione V3: Le dimensioni sociali e culturali della gestione forestale sostenibile
Risoluzione V4: La biodiversità forestale
Risoluzione V5: Cambiamenti climatici e gestione forestale sostenibile
Conferenza di Varsavia (2007):
• Risoluzione W1: Foreste, legno ed energia
6
La denominazione completa della convenzione è: Convention to Combat Desertification in Countries Experiencing Serious
Drought and/or Desertification, Particularly in Africa.
21
Davide Pettenella
• Risoluzione W2: Foreste e acqua
Un significato notevole, soprattutto nel monitoraggio e controllo del commercio internazionale del legname
tropicale, ha avuto la revisione dell'International Tropical Timber Agreement nel 1994 e l’attività
dell'International Tropical Timber Organisation (ITTO). Tra le azioni dell’ITTO merita segnalare la definizione
del "Year 2000 Objective", ovvero di un obiettivo formale consistente nella creazione delle condizioni di verifica
dell’origine del legname tropicale nonché della sostenibilità delle attività di estrazione e commercializzazione
dello stesso.
Nella Tabella 1 vengono presentate le principali iniziative internazionali, solo in parte richiamate in precedenza,
che hanno comportato interventi di regolamentazione del settore forestale. Dall’analisi dei contenuti dei trattati,
convenzioni e accordi definiti negli ultimi venti anni – e soprattutto negli anni '90 - emerge in forma evidente la
constatazione che il quadro normativo di riferimento per le politiche forestali è quanto mai articolato e
complesso. Tale quadro di norme è basato sia su strumenti giuridici vincolanti ("hard law") quali trattati,
protocolli, convenzioni, che - una volta ratificati – impegnano i Governi ad azioni coerenti (vd. Protocollo di
Kyoto), sia su strumenti "soft" che riflettono un consenso generalizzato, un accordo frutto di ampia
negoziazione, non giuridicamente vincolanti ma sostenuti da un notevole appoggio politico da parte delle
organizzazioni governative e, quindi, in genere caratterizzati da una effettiva volontà e capacità di una
implementazione operativa.
Un tentativo di lettura di sintesi delle diverse iniziative di regolamentazione può essere effettuata facendo
riferimento a tre campi di intervento, oggetto di un numero crescente di convenzioni, accordi, protocolli, ecc.
L’IFF ha utilizzato il termine “track” per definire i tre “sentieri di sviluppo” dell’attività normativa in sede
internazionale (vd. Tabella 2): quello connesso alla definizione di interventi di regolamentazione del mercato
internazionale, quello relativo alla promozione della gestione forestale sostenibile e, infine, quello delle misure
di tutela ambientale. E’ evidente che i tre gruppi di iniziative devono trovare un momento di coordinamento e
quindi un organismo superiore di negoziazione, quale potrebbe essere il già richiamato United Nations Forum on
Forests. E’ infatti del tutto illogico che, ad esempio, in sede di accordi del World Trade Organisation (WTO) si
tenti di eliminare le barriere non tariffarie al commercio internazionale, e tra queste vengano considerate le
iniziative di certificazione del legname, mentre in tutte le sedi in cui si affrontano i temi della gestione forestale
sostenibile si sostenga l’opportunità di una qualche forma di controllo e certificazione della provenienza del
legname.
Tabella 1 – Principali iniziative di regolamentazione del settore forestale definite in sede internazionale
CONVENZIONE/INIZIATIVA
SIGLA
SOFT/ RATIF. CONTENUTI PIÙ INNOVATIVI RELATIVI
HARD ITAL. AL SETTORE FORESTALE
Forest Principles (1992)
FP
Soft
Una serie di principi generali di riferimento
per le politiche forestali
Negotiated Proposals for Action
Soft
Predisposizione di più di 100 piani locali di
dell'IPF-IFF (1995-…)
azione grazie alla creazione di un ambito di
negoziazione internazionale specifico per il
settore forestale tra organismi governativi e
non e tra paesi del Sud e del Nord
Risoluzioni delle Conferenze PanS1,S2,
Soft
vd. Quadro 1
Europee per la Protezione delle
H1, H2,
Foreste
L1, L2
Convention on Conservation of
Hard Si
Mantenimento della biodiversità delle
European Wildlife and Natural
foreste europee, definizione di un network di
Habitats ("Convenzione di Berna")
aree protette. La convenzione è stata
(1979)
applicata nell'UE tramite le Dir. Habitat (Dir
43/92 e 62/97) e l'iniziativa Natura 2000, in
ambito europeo con l'iniziativa EMERALD
Convention on Wetlands of
Hard Si
Creazione di un network internazionale delle
International Importance
aree umide
("Convenzione di Ramsar") (1971)
Framework Convention on Climate FCCC e Hard Si
Inclusione delle foreste nei carbon budget;
Change (1992) e Protocollo di Kyoto KP
nuovi strumenti: AIJ e, successivamente, JIP
(1997)
e CDM
Convention on Biological Diversity CBD
Hard Si
Affermazione del principio di un equo
(1992)
utilizzo dei benefici derivanti dalla presenza
delle risorse genetiche; nuovi strumenti:
Clearing House Mechanism
22
Convention to Combat
Desertification (1992)
Convention on International Trade
in Endangered Species of Wild
Fauna and Flora (1973)
CCD
Hard
Si
Nuovi strumenti: Global Mechanism presso
IFAD; piani nazionali
Permesso di export e certificato di origine
per il commercio di specie legnose protette
(per esempio: Swietenia humilis, S.
mahagoni e S. macrophilla)
Attività dell'International Tropical Timber
Organisation; "Year 2000 Objective"; nuovi
strumenti: Bali Partnership Fund.
Incoraggiamento dei codes of conduct e dei
finanziamenti al settore di operatori privati;
impegno contro i tagli illegali, la corruzione
e il commercio irregolare del legname.
Identificazione e pubblicizzazione di
esperienze esemplari e dimostrative di
buona gestione forestale.
CITES
Hard
Si
International Tropical Timber
Agreement (1994)
ITTA
Hard
Si
G8 Action Programme on Forests
(1998)
G8APF
Soft
Si
International Model Forest Network
(su iniziativa del CAN) e
Demonstration Forest Management
Areas (CATIE) (anni ’90)
Long-Range Transboundary Air
Pollution Convention dell'UN ECE
(1979) e International Cooperative
Programme on Forests
-
Soft
-
LRTAP
e ICP
Forest
Hard
Si
Monitoraggio dell'impatto sulle foreste degli
inquinanti aerei.
Mediterranean Forest Action
Programme di Silva Mediterranea
MFAP
Soft
-
Protocollo Foreste della
Convenzione per la Protezione delle
Alpi (“Convenzione delle Alpi”)
(1991)
-
Hard
Si
Iniziative coordinate di gestione delle
foreste mediterranee. Stimolo alla visibilità e
responsabilità, riconoscimento della
ricchezza e fragilità delle foreste
mediterranee.
Le parti contraenti si impegnano affinché:
“siano adottati metodi di rinnovazione
forestale naturale; sia perseguita una
costituzione del patrimonio forestale ben
strutturata e graduata, con specie arboree
adatte al sito; sia impiegato un materiale di
riproduzione forestale autoctono; siano
evitate erosioni e compattazione del suolo,
mediante metodi di uso e di prelievo
rispettosi dell'ambiente”.
Tabella 2 - Quadro di sintesi delle iniziative di regolamentazione del settore forestale definite in sede
internazionale (Fonte: IFF Secretariat, 2000)
IL “SENTIERO” DEL
IL “SENTIERO” DELLA
IL “SENTIERO” DELLE
COMMERCIO
GESTIONE FORESTALE
PROBLEMATICHE AMBIENTALI
INTERNAZIONALE
SOSTENIBILE
WTO,
UNCED 1992:
CBD,
ITTA,
FCCC e KP,
• Principi Forestali,
CITES,
CCD,
• cap. 11 e altri di Agenda
FLEG,
LRTRAP e ICP Forest,
21,
altre iniziative
Convenz. di Berna e Ramsar,
IPF-IFF Negotiated Proposals
altre iniziative
for Action,
Risoluzioni Processo Paneuropeo,
G8APF, MFAP,
Protocollo Foreste,
altre iniziative
Campi prevalenti di intervento
Il commercio internazionale
La gestione, la conservazione e lo
La tutela delle risorse ambientali
sviluppo sostenibile delle foreste
3. Politiche di tutela del clima e interventi a favore della permanenza delle foreste
La ratifica del Protocollo di Kyoto (PK) ha comportato l’avvio di una serie di interventi per i paesi dell’Allegato
I del Protocollo, cioè per quei paesi che hanno assunto un obbligo di ridurre entro il 2008-2012 le proprie
23
Davide Pettenella
emissioni rispetto al 1990 (tra i quali l’Italia, che si è data un impegno di riduzione del 6,5% - Pettenella e
Zanchi, 2006).
Le misure per raggiungere gli impegni di riduzione delle emissioni di gas clima-alteranti riguardano in primis
il miglioramento dell'efficienza energetica in settori rilevanti dell'economia nazionale; la sviluppo e maggiore
utilizzazione di energia rinnovabile; l'adozione di misure per limitare le emissioni di gas ad effetto serra nel
settore dei trasporti, la limitazione e/o riduzione delle emissioni di metano attraverso il recupero e utilizzazione
del gas nel settore della gestione dei rifiuti, nonché nella produzione, il trasporto e la distribuzione di energia.
Il PK indica però anche altre strade, tra cui lo sviluppo di attività volte ad aumentare gli stock di carbonio negli
ecosistemi terrestri (Sedjo, 2006). Tali attività si concretizzano principalmente nelle nuove piantagioni su terreni
non coperti da vegetazione forestale (art. 3.3 del PK) e nel miglioramento degli stock delle foreste esistenti (art.
3.4).
Il PK definisce anche tre strumenti di mercato, noti come meccanismi flessibili, a cui i paesi dell’Allegato I
possono ricorrere per raggiungere i loro obiettivi nazionali di riduzione delle emissioni clima-alteranti in maniera
efficiente; essi sono:
•
Il Clean Development Mechanism (CDM), che consente ai paesi dell’Allegato I di investire in progetti
da realizzare nei paesi in via di sviluppo, in grado di ridurre le emissioni di gas-serra, ma anche di favorire lo
sviluppo tecnologico, economico e sociale dei paesi ospiti;
•
Il Joint Implementation (JI), che ammette la possibilità per i paesi dell’Allegato I di realizzare progetti
per la riduzione delle emissioni di gas-serra in un altro paese dello stesso gruppo e di utilizzare i crediti derivanti,
congiuntamente con il paese ospite;
•
l’Emissions Trading (ET), che riconosce la condizione di esercitare un commercio di crediti di
emissione tra i paesi dell’Allegato I, per esempio tra un paese che abbia conseguito una diminuzione delle
proprie emissioni di gas serra superiore al proprio obiettivo e un paese che viceversa non sia stato in grado di
rispettare i propri impegni di riduzione delle emissioni di gas-serra. Questo meccanismo flessibile è stato
implementato nell’Unione Europea (UE) con la creazione dell’Emission Trading System (ETS), un mercato nel
quale tutte le imprese dell’UE che operano in alcuni settori economici energy intensive devono rispettare degli
obblighi di emissione annualmente definiti; per rispettare questi obblighi le imprese possono acquistare crediti di
emissione disponibili sul mercato. L’ETS è attualmente il più grande mercato mondiale di scambio di quote di
carbonio (Capoor e Ambrosi, 2007), anche se – per l’impostazione ad esso dato dalla Commissione Europea – in
questo mercato non possono essere venduti crediti di emissione provenienti dalla realizzazione di attività nel
settore agricolo e forestale in Europa.
L’utilizzo di tali strumenti, come l’attività di rendicontazione dei risultati nazionali relativi all’attuazione del PK,
sono in Italia di competenza delle autorità centrali dello Stato. In altri termini non si è ritenuto finora opportuno
trasferire alle Regioni e alle Province Autonome gli obblighi formali di implementazione del PK assunti dal
Governo italiano in sede internazionale (Ciccarese et al., 2006).
Peraltro da molti anni diverse amministrazioni pubbliche, imprese (vd. ad esempio le iniziative di The Climate
Group www.theclimategroup.org) e perfino singoli individui (Quadro 3 e Figura 1) hanno sentito la necessità di
effettuare investimenti di tipo volontario per ridurre o annullare le proprie emissioni (vd. come caso limite il
Quadro 4 che presenta il primo e per ora unico caso di uno Stato “Carbon neutral”). Tali scelte sono legate a
motivazioni ideali, ma anche a considerazioni pragmatiche connesse ai risparmi e alla maggior competitività
delle imprese con bassi livelli di emissioni e anche all’utilizzo di tecniche di green marketing volte a migliorare
l’immagine dell’organizzazione nel mercato.
Gli investimenti volontari si basano su iniziative che interessano direttamente le attività delle singole
organizzazioni (ad esempio un piano di razionalizzazione della mobilità dei dipendenti di una azienda) o su
interventi compensativi (Carbon offset) esterni al contesto in cui operano le organizzazioni, quali ad esempio il
finanziamento della realizzazione in un paese in via di sviluppo di un piano di rimboschimento o di una centrale
eolica.
Sugli investimenti di carattere volontario, e in particolare su quelli che interessano le risorse forestali e che
possono influenzarne la conservazione e permanenza, si concentra l’attenzione nel seguito del testo.
Quadro 3 – Una tesi di laurea “a impatto zero”. Presso l’Università di Padova nell’anno accademico 2006-07
è stata realizzata da Elisa Negrini una tesi di laurea con il titolo “Bilancio delle emissioni di gas effetto serra
delle
facoltà
di
Agraria
e
Medicina
Veterinaria
Polo
di
Agripolis,
Legnaro”
(www.tesaf.unipd.it/pettenella/tesi/Negrini.htm ). La tesista ha conteggiato tutte le emissioni di gas di serra
collegate alla conduzione della tesi, ivi compresi i trasferimenti dalla sua residenza ai diversi siti dove sono stati
raccolti i dati. Come evidenziato nella figura 1, le emissioni per la realizzazione della tesi (pari a 920 kg di CO2)
sono state compensate con una piantagione forestale in Costa Rica di 1.187 mq. Il costo dell’investimento
compensativo è stato di 60€, per cui i costi unitari sono risultati pari a 65,2 €/t CO2 e di 505 €/ha rimboschito.
Quadro 4 – L’investimento realizzato in Ungheria per compensare le emissioni della Santa Sede. Due
aziende, la statunitense Planktos e l'ungherese Klimafa, su base volontaria hanno deciso di compensare le
24
emissioni di anidride carbonica prodotte dal Vaticano nel 2007. Grazie ad un inventario si è arrivati a stimare tali
emissioni in 7.500 tonnellate di anidride carbonica all’anno. Verrà quindi realizzata una piantagione in Ungheria
di 15 ettari in grado di assorbire, nel corso della sua crescita secolare, i gas serra emessi dal Vaticano nel 2007.
3.1 Le caratteristiche degli investimenti forestali
volontari
La realizzazione di interventi di carattere volontario
consente ai diversi investitori pubblici e privati una
maggior flessibilità e una maggior gamma di
interventi non essendo necessariamente soggetti alle
limitazioni e regole imposte dal PK. Ad esempio
possono essere programmati interventi di riduzione
delle emissioni legate ai fenomeni della
deforestazione e della degradazione delle foreste
(nel gergo Reducing Emisssions from Deforestation
and Degradation of Forest, REDD), che sono
all'origine del 20% circa delle emissioni globali di
gas-serra. Questa linea di intervento non è
attualmente prevista nel PK, anche se, nell’ultima
Conferenza delle Parti relativa all’attuazione del PK
tenutasi a Bali, è stato deciso di sperimentare i
diversi approcci metodologici per la riduzione della
deforestazione e del degrado delle foreste nei
prossimi due anni, in vista della revisione del PK
che verrà effettuata nella prossima Conferenza delle
Parti di Copenaghen.
Va comunque tenuto presente che la realizzazione
di interventi forestali nei paesi in via di sviluppo
nella logica di un miglioramento effettivo delle
condizioni climatiche, della tutela della biodiversità
e delle condizioni socio-economiche locali non è
priva di problemi (Peskett et al., 2006). Gli aspetti
Figura 1 – Dichiarazione relativa ad un
problematici principali sono i seguenti:
investimento compensativo nel settore forestale
- il rispetto del criterio dell’addizionalità
degli investimenti realizzati in relazione alle condizioni ordinarie di gestione, cioè allo scenario
“business as usual”; perché si tratti di investimenti compensativi volti effettivamente a migliorare le
condizioni climatiche le attività realizzate devono infatti essere caratterizzate da una intenzionalità
esplicita: non avrebbe senso, ad esempio, acquisire crediti di carbonio provenienti da un foresta già in
fase di abbandono, il cui stock aumenta indipendentemente dalla realizzazione di un investimento
compensativo;
- la verifica della permanenza degli effetti: un altro aspetto che non va trascurato nei casi di intervento
sulle risorse forestali è infatti legato al verificarsi di fenomeni, intenzionali o non voluti, che
determinano il ritorno in atmosfera del carbonio fissato (ad esempio: incendi, schianti, danni causati da
attacchi di insetti, ecc.);
- la necessità di evitare effetti collaterali di segno opposto a quelli dell’investimento compensativo
realizzato e da questo dipendenti, il problema definito del “leakage”; ad esempio, la messa in protezione
di una foresta da fenomeni di deforestazione illegale non potrebbe essere considerata un investimento
compensativo corretto se fosse accompagnata dal trasferimento su altre aree forestali limitrofe delle
medesime attività illegali.
Altri problemi che si possono incontrare nel mettere in atto investimenti compensativi nel settore forestale sono
legati alla complessità tecnica e al costo economico delle attività di inventariazione e monitoraggio e dal fatto
che, anche per questi elementi di complessità, c’è il rischio di privilegiare gli interventi su grandi superfici dove
è facile realizzare economie di scala negli investimenti e nella valutazione dei relativi effetti, “spiazzando”
quindi gli interventi su piccola scala. Su questo aspetto problematico (grandi progetti con ottimi effetti di
immagine ed economie di gestione rispetto a micro-realizzazioni con effetti più diffusi, maggiore controllo
sociale delle popolazioni interessate ma costi di gestione e monitoraggio più alti) va trovato un corretto
equilibrio. Un ultimo, ma certamente non secondario, problema è quello della equa ripartizione dei benefici degli
interventi (Grieg-Gran et al., 2005): c’è infatti un’ampia evidenza empirica che in molti progetti larga parte dei
pagamenti non vada ai gestori delle risorse ma alle diverse figure di intermediari coinvolti nelle transazioni.
3.2 Le modalità di funzionamento dei mercati degli investimenti volontari
25
Davide Pettenella
Gli investimenti compensativi sono favoriti dalla presenza di intermediari o agenzie di servizio7 (vd. figura 2)
che offrono un portafoglio di possibili interventi, mettendo in relazione le organizzazioni che offrono progetti di
investimento e quelle che intendono acquistare i crediti da questi derivanti. La tabella 3 riporta alcune delle
principali organizzazioni a livello internazionale che fanno da intermediazione tra i soggetti che propongono
investimenti e quelli che esprimono una domanda di interventi compensativi. Come evidenziato nella tabella, le
agenzie di intermediazione si specializzano in una serie ristretta di tipologie di investimenti, altre operano su una
gamma molto ampia.
Tabella 3 - Iniziative volontarie di compensazione delle emissioni di carbonio
Organizzazione Prezzo NonTipo di progetto
Possibilità
Attività compensabili
responsabile
medio profit
scelta del
dell’offerta di (US$/
progetto
investimenti
ton
compensativi
CO2)
Autobonfund.org $4,30- Sì Rinnovabili, Riforestazione/
Sì
Gestione domestica,
USA
5,50
afforestazione
Auto, Aereo, Eventi,
Attività economiche
Sistema di
garanzia
Green-e,
Chicago
Climate
Exchange,
Environmental
Resources
TrUSAt
Chicago
Climate
Exchange,
Environmental
Resources
TrUSAt
n.d.
e-BlueHorizons
USA
$5,00
No Rinnovabili,
Riforestazione/afforestazion
e
No
Gestione domestica,
Auto, Aereo
Greenfleet
Australia
$7,007,50
Sì Riforestazione/afforestazion
e
No
Gestione domestica,
Auto, Aereo
DrivingGreen
Irlanda
Terrapass
USA
$8,00
No Rinnovabili
No
Auto
n.d.
$8,8011,00
No Rinnovabili,
No
$10,00
Sì Rinnovabili
No
Calco9li esterni
Green-e,
Chicago
Climate
Exchange
n.d.
No Riforestazione/afforestazion
e
$13,00- No Rinnovabili, ,
27,00
Riforestazione/afforestazion
e
No
Gestione domestica,
Auto, Aereo, Babies
Gestione domestica,
n.d.
Native Energy
USA
$13,20
No Rinnovabili
Sì
Green-e
Climate Friendly
Australia
$16,00- No Rinnovabili
19,00
No
Gestione domestica,
Gestione domestica,
Auto, Aereo, Attività
economiche
Solar Electric
Light Fund
USA
Autobon Clear
Regno Unito
Autobon Neutral
Company
Regno Unito
7
$17,00
Sì
n.d.
Office of the
Renewable
Energy
Regulator,
NSW
Government,
Ernst &
Young.
In Italia, anche nell’offerta di investimenti nel settore forestale, operano tra gli altri AzzeroCO2 (www.azzeroco2.it) e
LifeGate (www.impattozero.it). AzzeroCO2, nella proposta di investimenti compensativi legati ad attività forestali, si collega
a Carbon Neutral (www.carbonneutral.com), una organizzazione operante ormai da diversi anni specificatamente nel settore
forestale.
26
SUSAtainable
travel International
USA, Svizzera
Trees for Life
Regno Unito
Grow a Forest
Regno Unito
$18,00
Sì Rinnovabili
No
Aereo, Auto, Gestione
domestica, Hotel
See
Myclimate
$20,00
appr.
$22,00
& Up
Sì Riforestazione/afforestazion
e
No Riforestazione/afforestazion
e
No
n.d.
Bonneville
Environmental
Foundation
USA
Myclimate
Svizzera
$29,00
Sì Rinnovabili
No
Auto, Aereo, Trasporti
pubblici
Aereo, Auto, Gestione
domestica, Attività
economiche
Gestione domestica,
Aereo, Attività
economiche, Eventi
$30,00
Sì Rinnovabili
No
No
Aereo, Eventi, Attività
economiche
n.d.
Green-e
DeSìgnated
Operational
Entity
Note: "n.d." significa che non si è potuta determinare la presenza di un sistema di garanzia di parte terza. Il
progetto può, comunque, essere sottoposto a verifica.
Prezzo medio: i prezzi cambiano e i tassi di scambio sono oggetto di variazione. Quelli riportati si riferiscono al
luglio 2006.
Da: http://www.ecobusinesslinks.com (rielaborato)
Offerta = soggetti che vogliono realizzare
investimenti di C offset
Consulenti/
progettisti
Standard
CCX
OTC
Consulenti/
progettisti
Verificatori o
certificatori
Intermediari
Mercati di
scambio
Marchi
Domanda = soggetti che vogliono
compensare le proprie emissioni
Figura 2: Schema dell’organizzazione del mercato degli investimenti volontari
I crediti di carbonio creati per il mercato degli interventi volontari sono generalmente chiamati VERs (Verified
Emission Reductions). Una tonnellata di emissioni di CO2 ridotta genera un VER. Il mercato dei VERs è
cresciuto da un valore di circa 4 milioni di VERs commercializzate nel 2004 a circa 100 milioni nel 2007. In
base al rapporto State of the Voluntary Carbon Markets 2008 (Bayon et al., 2008), la crescita degli scambi ha
interessato negli ultimi anni, soprattutto i paesi asiatici, gli Stati Uniti, l’Australia e la Nuova Zelanda. Il mercato
più ampio dove si scambiano i VERs è il Chicago Climate Exchage (CCX) nel quale dal 2003, anno di lancio del
CCX, sono stati scambiati circa 15 milioni di quote (vd. figura 3).
Il prezzo dei VERs, date le procedure più snelle di approvazione, verifica e monitoraggio dei progetti da
iniziative volontarie, è risultato sempre inferiore rispetto a quello delle quote relative ai mercati ufficiali (l’ETS
dell’UE). Attualmente (maggio 2008), il prezzo di una quota per interventi compensativi nel mercato volontario
(CCX) è di circa 7 $ (4,5 €), mentre quello delle quote nell’ETS è di 35 $ (22,1 €).
27
Davide Pettenella
Nel 2007 il mercato dei VERs ha interessato 65 milioni tonnellate (Mt) di CO2 equivalente, una crescita molto
significativa rispetto ai 25 Mt del 2006. Dei 65 Mt, 23 Mt sono state scambiate nel Chicago Climate Exchange e
42 Mt sono state commercializzate in mercati bilaterali o non standardizzati (definiti OTC “over the counter”).
La gamma dei prezzi dei VERs è molto ampia, anche per la frammentazione dei luoghi e dei soggetti delle
contrattazioni: da 1,8 a 300 $/t CO2. Insieme ai progetti relativi alle rinnovabili, quelli connessi a piantagioni
forestali hanno espresso i prezzi più alti: da 6,8 $ per interventi con specie autoctone a 8,2 $ per le monoculture,
spesso realizzate con specie esotiche (i prezzi sono influenzati al costo di acquisizione dei terreni e di
piantagione). Gli interventi di contenimento della deforestazione (“Avoided deforestation”) hanno avuto prezzi
medi di 4,8 $, mentre quelli di sola conservazione del suolo di 3,9 $/t.
Figura 3: Andamento dei prezzi e dei volumi di VERs nel Chicago Climate Exchage
(fonte: http://www.chicagoclimatex.com/market/data/summary.jsf)
Il rapporto citato evidenzia la recente diminuzione relativa degli investimenti compensativi nel settore forestale:
questi rappresentavano il 37% dei VERs nel 2006, mentre nel 2007 hanno coperto il 18% delle quote scambiate.
Le ragioni di questa diminuzione sono legate ai requisiti più rigorosi imposti ai progetti forestali che si traducono
in difficoltà più ampie nel trovare siti dove tutti gli impatti ambientali e sociali siano positivi e correttamente
valutati. In particolare il rispetto della “addizionalità” sembra ostacolare molto lo sviluppo dei progetti di
riduzione della deforestazione (REDD) nei paesi in via di sviluppo.
In effetti la ripartizione dei progetti vede, al 2007, prevalere quelli legati a nuove piantagioni (42% con specie
native e 13% di monoculture), mentre i progetti REDD coprono il 28% del mercato; la parte residua interessa
interventi di tutela dei suoli e di altre formazioni naturali, quali le aree umide.
Lo sviluppo recente di standard e sistemi di verifica indipendente, a cui si accenna nel seguito, può contribuire a
rilanciare gli investimenti compensativi in campo forestale. Secondo gli analisti, la possibile inclusione dei
progetti REDD nell’attuazione del PK porterà ad una maggiore attenzione a questi progetti anche nel mercato
volontario.
3.3 I sistemi di controllo dei progetti volontari
Il differenziale dei prezzi tra il mercato ETS e quello dei VERs è legato non solo alla diversa tipologia degli
investimenti ma anche, come accennato, ai sistemi di garanzia che vengono offerti sull’effettiva capacità di
fissazione di carbonio degli investimenti realizzabili e realizzati (Peskett et al., 2006). La veridicità delle
dichiarazioni può essere garantita sulla base di attestazioni dell’agenzia di intermediazione, di organismi esterni
o in base a certificazioni di enti terzi indipendenti (anche accreditati).
Secondo una stima prudenziale di Bayon et al. (2008), circa la metà delle transazioni che hanno interessato i
VERs nel settore delle risorse agricole-forestali si sono basate sull’impiego di standard indipendenti. Gli
standard impiegati, ovviamente non solo nel campo degli investimenti volontari, sono classificabili in tre gruppi:
- gli standard generici per la stima degli effetti sul ciclo del carbonio e, in alcuni casi, degli impatti
ambientali e sociali degli investimenti; il Gold Standard (Ecofys, 2006), il Voluntary Carbon Standard,
lo standard per i CDM, il CCX e il VER+ sono quelli più impiegati;
28
gli standard di buona gestione forestale in base a criteri ambientali, sociali ed economici; la Banca
Mondiale per i progetti forestali CDM richiede la certificazione in base allo standard del Forest
Stewardship Council (FSC), che è stato impiegato anche negli investimenti volontari; per i progetti
forestali offerti nel Chicago Climate Exchange è stata recentemente ammessa anche la certificazione in
base allo standard del Programme for the Endorsement of Forest Certification (PEFC) Schemes;
entrambi gli standard, tuttavia, non contemplano una procedura per la rigorosa valutazione della
capacità di fissazione di carbonio;
- anche per questa ragione sono stati sviluppati di recente, negli USA e in Europa, due standard specifici
relativi ai progetti forestali compensativi (il Climate, Community and Biodiversity Standard e il
CabonFix Standard).
Già almeno un paio di grandi organizzazioni internazionali di certificazione hanno predisposto programmi di
certificazione indipendente degli interventi compensativi forestali, soprattutto nella prospettiva di un controllo
degli investimenti realizzati con i meccanismi flessibili del Protocollo di Kyoto. Evidentemente tali attività di
certificazione offrono maggiori tutele agli investitori, ma alzano ulteriormente i costi amministrativi degli
investimenti.
In effetti per le attività di compensazione realizzate nel settore forestale, le procedure ufficiali di approvazione e
monitoraggio sono molto complesse, tanto che nel mercato dei CDM (UNDP, 2006) è stato finora approvato un
solo progetto forestale (una piantagione realizzata in Cina con una capacità di fissazione di 25.795 t CO2/anno vd. http://cdm.unfccc.int/Projects/registered.html).
-
Considerazioni finali
Alla luce dell’analisi effettuata in queste pagine è possibile affermare che la definizione ufficiale di foresta
adottata in sede di inventari mondiali dalla FAO è talmente ampia da non cogliere i problemi principali di quella
fase di evoluzione del settore, che sono quelli del degrado e progressivo impoverimento delle formazioni. Dalle
statistiche internazionali emerge una condizione di apparente permanenza di una ampia copertura forestale, parte
della quale, agli occhi di un non esperto, potrebbe essere meglio definita come terreni nudi con sporadica
presenza di piante arboree.
In questo contesto gli strumenti di regolamentazione internazionale cogenti (“hard law”, tra i quali in primis il
Protocollo di Kyoto) non consentono di supportare efficacemente la permanenza delle foreste andando a
privilegiare le azioni volte al rimboschimento.
Gli investimenti volontari che, sull’onda delle preoccupazioni legate ai cambiamenti climatici, affrontano il
problema del REDD e della non permanenza, sono caratterizzati ancora da alti i costi di transazione necessari per
garantire le condizioni di corretta esecuzione degli interventi.
Lo sviluppo del mercato degli investimenti compensativi in campo forestale è peraltro un positivo elemento di
novità nell’organizzazione del settore in quanto rende operativo quel principio “chi fornisce benefici ambientali
viene remunerato” (“Provider gets”) complementare a quello universalmente accettato del “chi inquina paga”
(“Polluters pays”). Compensare gli interventi addizionali di fissazione del Carbonio, soprattutto quando questi
sono realizzati in territori economicamente marginali, non è solo un principio di efficiente gestione del mercato,
ma anche una scelta eticamente corretta, sempre che gli investimenti considerino anche requisiti di tutela
ambientale e sociale (Peskett et al., 2007).
Bibliografia
Bayon R., A.Shapiro, S.Zwick, 2008. Forging a Frontier: State of the Voluntary Carbon Markets 2008.
Ecosystem Marketplace, New Carbon Finance; Washington, New York (in: http://ecosystemmarketplace.com )
Bennett E.L., J.G.Robinson, 2000. Hunting of wildlife in tropical forests. Implications for biodiversity and forest
people. Environment Department Papers (76). World Bank, Washington D.C.
Buttoud G., 2001. Gérer les forêts du sud. L’Harmattan, Paris.
Capoor K., P.Ambrosi, 2007. State and Trends of the Carbon Market 2007. World Bank, Washington (in:
http://carbonfinance.org/docs/Carbon_Trends_2007-_FINAL_-_May_2.pdf)
Ciccarese L., D.Pettenella, G.Zanchi, 2006. Il settore primario e la riduzione delle emissioni di gas ad effetto
serra. Tra strumenti diretti di compensazione e politiche generiche di sostegno del settore. Politica Agricola
Internazionale (5), pp. 27-48.
Commissione Europea, 2003. L’applicazione delle normative, la governance e il commercio nel settore forestale
(FLEGT). Proposta di un piano d’azione dell’Unione Europea. Comunicazione della Commissione al Consiglio e
al Parlamento Europeo. Documento COM (2003) 251 definitivo. Bruxelles, 21.05.2003.
Ecofys, 2006. The Gold Standard: Voluntary emission reductions (VERs). Manual for Project Developers,
Version 5. Ecofys. http://www.cdmgoldstandard.org/
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Davide Pettenella
Food and Agriculture Organization-International Tropical Timber Organization, 2005. Best practices for
improving law compliance in the forestry sector. FAO, Rome.
Global Witness, 2002. The Logs of war. The timber trade and armed conflict. GW, London.
Grieg-Gran M., I.Porras, S.Wunder, 2005. How can Market Mechanisms for Forest Environmental Services
Help the Poor? World Development, 33 ( 9), pp. 1511–1527
Humphreys D., 1996. Forest Politics: The Evolution of International Cooperation, Earthscan.
Mery G., R.Alfaro, M.Kanninen, M.Lobovikov (eds), 2005. Forest in the global balance. Changing paradigms.
IUFRO World Series, vol. 17. IUFRO, Vienna.
Palo M., 1990. Deforestation and development in the third world: roles of system causality and population. In:
Palo M., G.Mery (eds), 1990. Deforestation or development in the Third World? Volume III.
Metsäntutkimuslaitoksen Tiedonantoja 349. Division of Social Economics of Forestry.
Palo M., J.Salmi (eds), 1987. Deforestation or development in the Third World? - Metsäntutkimuslaitoksen
Tiedonantoja 272. Division of Social Economics of Forestry, The Finnish Forest Research Institute. Helsinki.
Peskett L., C.Luttrell, D.Brown, 2006. Making voluntary carbon markets work better for the poor: the case of
forestry off-sets. Overseas Development Institute, Forestry Briefing (11).
Peskett L., C.Luttrell, M.Iwata, 2007. Can standards for voluntary carbon offsets ensure development benefits?
Overseas Development Institute, Forestry Briefing (13).
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settore forestale, in R. Pilli, T.Anfodillo, E.Dalla Valle (eds), Stima del carbonio in foresta: metodologie e aspetti
normativi, Atti del 42° corso di Cultura in Ecologia, Università di Padova: 1-10 (in
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Scotland N., S.Ludwig, 2002. Deforestation, the timber trade and illegal logging. EC Workshop on Forest Law,
Enforcement, Governance and Trade, Brussels, April 22-24, 2002, 9 p.
Sedjo R., 2006. Forest and Biological Carbon Sinks after Kyoto. Resources for the Future, Washington.
Sist P., R. Fimbel, D. Sheil, R. Nasi, M.H. Chevallier, 2003. Towards Sustainable Management of Mixed
Dipterocarp Forests of Southeast Asia: Moving Beyond Minimum Diameter Cutting Limits. Environmental
Conservation, 30 (4).
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literature review and annotated bibliography. International Institute for Environment and Development, London.
Transparency International, 2007. Corruption Perceptions Index 2007 (in
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Programme, New York.
World Bank, 2002. Revised Forest Strategy. World Bank, Washington, DC.
30
Allegato
Le definizioni di Forest, Other wooded, Deforestation, Forest Degradation e Plantations
Fonti: Forest Resources Assessment
11/CP.7 UNFCCC, 2001)
della FAO, www.fao.org/docrep/009/j9345e/j9345e07.htm; Decision
Forest
Land with tree crown cover (or equivalent stocking level) of more than 10 percent and area of more than 0.5
hectares (ha). The trees should be able to reach a minimum height of 5 meters (m) at maturity in situ. May
consist either of closed forest formations where trees of various storeys and undergrowth cover a high proportion
of the ground; or open forest formations with a continuous vegetation cover in which tree crown cover exceeds
10 percent. Young natural stands and all plantations established for forestry purposes which have yet to reach a
crown density of 10 percent or tree height of 5 m are included under forest, as are areas normally forming part of
the forest area which are temporarily unstocked as a result of human intervention or natural causes but which are
expected to revert to forest.
Includes: forest nurseries and seed orchards that constitute an integral part of the forest; forest roads, cleared
tracts, firebreaks and other small open areas; forest in national parks, nature reserves and other protected areas
such as those of specific scientific, historical, cultural or spiritual interest; windbreaks and shelterbelts of trees
with an area of more than 0.5 ha and width of more than 20 m; plantations primarily used for forestry purposes,
including rubberwood plantations and cork oak stands.
Other wooded land
Land either with a crown cover (or equivalent stocking level) of 5-10 percent of trees able to reach a height of 5
m at maturity in situ; or a crown cover (or equivalent stocking level) of more than 10 percent of trees not able to
reach a height of 5 m at maturity in situ (e.g. dwarf or stunted trees); or with shrub or bush cover of more than 10
percent.
Deforestation
UNFCCC: The direct human-induced conversion of forested land to non-forested land.
FAO: The conversion of forest to another land use or the long-term reduction of the tree canopy cover below the
minimum 10 percent threshold
Deforestation implies the long-term or permanent loss of forest cover and implies transformation into
another land use. Such a loss can only be caused and maintained by a continued human-induced or
natural perturbation.
It includes areas of forest converted to agriculture, pasture, water reservoirs and urban areas.
The term specifically excludes areas where the trees have been removed as a result of harvesting or
logging, and where the forest is expected to regenerate naturally or with the aid of silvicultural
measures. Unless logging is followed by the clearing of the remaining logged-over forest for the
introduction of alternative land uses, or the maintenance of the clearings through continued disturbance,
forests commonly regenerate, although often to a different, secondary condition. In areas of shifting
agriculture, forest, forest fallow and agricultural lands appear in a dynamic pattern where deforestation
and the return of forest occur frequently in small patches. To simplify reporting of such areas, the net
change over a larger area is typically used.
Deforestation also includes areas where, for example, the impact of disturbance, over-utilization or
changing environmental conditions affects the forest to an extent that it cannot sustain a tree cover
above the 10 percent threshold.
Forest degradation
Takes different forms, particularly in open forest formations, deriving mainly from human activities such as
over-grazing, over-exploitation (for firewood or timber), repeated fires, or due to attacks by insects, diseases,
plant parasites or other natural sources such as cyclones. In most cases, degradation does not show as a decrease
in the area of woody vegetation but rather as a gradual reduction of biomass, changes in species composition and
soil degradation. Unsustainable logging practices can contribute to degradation if the extraction of mature trees
is not accompanied with their regeneration or if the use of heavy machinery causes soil compaction or loss of
productive forest area.
Plantations
Afforestation: artificial establishment of forest on lands which previously did not carry forest within
living memory.
Reforestation: Artificial establishment of forest on lands which carried forest before.
31
32
FOREST FIRES AND AIR POLLUTION
Ana Isabel MIRANDA, C. BORREGO
University of Aveiro - CESAM, Department of Environment and Planning
[email protected]
FOREST FIRES AND AIR POLLUTION
1 Introduction
A forest fire is a large-scale natural combustion process consuming various types, sizes and ages of botanical
specimen growing outdoors in a defined geographical area. Although wildland fires are an integral part of
ecosystems management and are essential to maintain functional ecosystems (Sandberg et al. 2002), their
dimensions can give rise to disastrous results and due to the frequency of occurrence and the magnitude of
effects on the environment, health, economy and security, forest fires have increasingly become a major subject
of concern for decision-makers, firefighters, researchers and citizens in general. Within their consequences, it is
the emission of various environmentally significant gases and solid particulate matter to the atmosphere that
interfere with local, regional and global phenomena in the biosphere.
Smoke from forest fires contains important amounts of carbon dioxide (CO2), carbon monoxide (CO), methane
(CH4), nitrogen oxides (NOx), ammonia (NH3), particulate matter (PM), particles with a mean diameter less than
2.5 µm (PM2.5) and particles with a mean diameter less than 10 µm (PM10), non-methane hydrocarbons (NMHC)
and other chemical compounds. These air pollutants can cause serious consequences to local and regional air
quality by reducing visibility, contributing to smog and impairing air quality in general, thus threatening human
health and ecosystems. CO, CH4, NMHC, and NOx are chemically active gases that strongly influence the
local/regional concentrations of the major atmospheric oxidants (ozone (O3) and the hydroxyl radical OH), and
high levels of tropospheric ozone can occur at great distances from emission sources (Crutzen and Andreae,
1990; Crutzen and Carmichael, 1993). Production of aerosols is also very important, giving rise to local
pollution, and affecting the radiation budget of the Earth and, hence, impacting global climate. Globally, fires are
a significant contributor of CO2 and other greenhouse gases to the atmosphere. Fires account for approximately
one fifth of the total global emission of CO2 (Sandberg et al., 2002). Even taking into account the notion that
fires in temperate ecosystems are minor contributors compared to biomass burning in savannas, boreal and
tropical forests, contribution to total CO2 equivalent emissions produced during forest fires can reach 7% if the
annual area burned exceeds 100,000 ha (Miranda et al., 1994a).
Moreover, smoke pollution due to forest fire events can represent an important public health issue to the
community, particularly for personnel involved in firefighting operations (Brustet et al., 1991; Ward et al., 1993;
Miranda et al., 1994b; Reinhardt et al., 2001; Miranda et al., 2005a; Valente et al., 2007).
Severe air pollution episodes caused by fires in Amazonia (Brazil), Indonesia and Philippines in 1997/98 and,
more recently, in Australia and Russia, have drawn worldwide attention to the problem of air quality due to
forest fires. Increasingly, smoke pollution due to wildland fires is considered an important health issue with
major risks for the population and the environment. The World Health Organization (WHO) has even provided
guidelines for forest fire episodic events to protect the public from adverse health effects (WHO, 1999). This
concern also applies to prescribed fires, especially in Australia and North America where this land management
technique is frequently used.
The main purpose of this manuscript is to provide an overview of the relation between forest fires and air
pollution, taking into account emissions, and pollutants dispersion and chemistry. Also, several examples will be
given to illustrate experimental and modelling case studies.
2 Atmospheric Emissions from Forest Fires
Forest fire atmospheric emissions depend on multiple and interdependent factors like forest fuels characteristics,
burning efficiency, burning phase, fire type, meteorology and geographical location.
- Fuel type and load are one of the most important factors affecting fire emissions. Variations in fuel
characteristics and consumption may contribute to 30% of the uncertainties in estimates of wildfires emissions
(Peterson, 1987; Peterson and Sandberg, 1988). This is a critical factor when describing forest fuels in the SouthEuropean forests, because available fuel mass depends on the location, fuel type and time of the year.
- Burning efficiency is also a significant fire emissions factor, which is usually defined as the ratio of carbon
released as CO2 to total carbon present in the fuel. For convenience, the modified combustion efficiency can also
be used; meaning the ratio of carbon released as CO2 by the sum of CO2 and CO (Ward, 1999) and is directly
related to the vegetation type and its moisture content. In laboratory and field experiments, the burning
efficiency can be expressed as the fraction burned related to the total biomass available.
- Emissions estimates for biomass burning distinguish different combustion phases. The durability or
predominance of burning phases depends on fuel type/mixture, moisture content and atmospheric conditions. For
biomass burning, it can be considered the following phases: pre-ignition, flaming, smoldering and glowing.
During the flaming phase, the most emitted compounds are CO2 and water vapor and, in less quantity, NOx,
34
Forest fires and air pollution
sulphur dioxide (SO2), nitrogen (N2) and particles with high carbon content. Emissions more oxidized are
predominant as a result of higher burning efficiency. In the smoldering phase, emissions partially oxidized or
reduced are predominant, namely CO and others like CH4, NMHC and polycyclic aromatic hydrocarbons (PAH),
NH3, sulphur compounds (COS, DMS e DMDS) and particles with low black carbon content. CO is the main
compound emitted during this phase.
- Concerning the type of fire, heading fires have high propagation speed and fuel consumption rate, and
therefore low burning efficiency and less oxidized emissions. Backing fires have low propagation speed and
long residence time, and therefore high burning efficiency and more oxidized emissions.
- Meteorological parameters as air temperature, humidity, precipitation, wind direction and intensity are other
factors affecting the moisture content of fine fuels and soil duff. Fuels dryness increases flammability and flame
propagation speed. Atmospheric stability has also a significant role in fire behaviour in the initial combustion
phases. Rain and hail are important meteorological elements to be considered in fire extinction, but can also
contribute to the increase of the available biomass fuel after winter and spring.
2.1 Emissions modelling
Emissions from forest fires can be estimated using models, e.g. FOFEM (Reinhardt et al.,1997), EPM (Sandberg
and Peterson, 1984), CONSUME (Ottmar et al., 2002), AIR FIRE (Trozzi et al., 2002) or EMISPREAD
(Miranda et al., 2005b). They are frequently based on a simplified methodology, which include the emission
factors, the burning efficiency, the fuel loads and the burned area. Hence, emissions can be estimated through the
following simple expressions:
Ei = A × B × β × FEi
or
Ei = A × FEi
in which
A – available fuel area (m2)
B – fuel load (kg.m-2)
β – burning efficiency (as fraction of biomass burned)
FEi – emission factor for pollutant i (g.kg-1 or kg.ha-1)
Ei – emissions for pollutant i (g)
Emission factors are defined as mass of pollutant emitted per mass of burned fuel mass (g.kg-1) or burned area
(g or kg.ha-1). A great variety of emission factors can be found in the literature, dealing with different fire types,
burning phases and fuel types. Emission factors are not easily derived, because several factors have to be
considered like fuel consumption, vegetation type, burned area and fire conditions. Therefore, there is a
significant difference between emission factors from experimental and prescribed fires compared to emission
factors from wildfires. The use of emission factors from the other type of fires to predict emissions in wildfires
may result in great error. Determination of emission factors contributes to 16% in estimative errors for forest fire
emissions (Peterson, 1987; Peterson and Sandberg, 1988).
Miranda (2004) presented a selection of emission factors for South-European forest fires. This review and
summarizing work was further updated and expanded, taking into account more recent works and more
pollutants. It was possible to compile in Table I the emission factors for the following air pollutants: CO2, CO,
CH4, total particulate matter (TP), PM2.5, PM10, NMHC, NOx, NH3 and SO2, for different type of fires and
burning phases.
Table I. Averaged emission factors (g.kg-1) for South-European forests (Miranda et al., 2005b).
Emission factors (g.kg-1)
Resinous w/
Pollutants Burning Herbaceou Brushwood
Eucalyptu Deciduou
phase/
Resinous brushwood
s
s
s
s
fire type
understory
CO2
CO
F/B
S/H
G
F/B
S/H
1450
1414
1418
78
106
1509
1426
1477
88
95
1704
1440
1627
55
180
1562
1421
1487
64
155
1530
1327
1414
57
161
1537
1293
1393
46
183
35
Emission factors (g.kg-1)
Resinous w/
Pollutants Burning Herbaceou Brushwood
Eucalyptu Deciduou
phase/
Resinous brushwood
s
s
s
s
fire type
understory
CH4
NMHC
TP
PM2,5
PM10
NOx
SO2
NH3
G
F/B
S/H
G
F/B
S/H
G
F/B
S/H
G
F/B
S/H
G
F/B
S/H
G
F/B
S/H
G
F/B
S/H
G
F/B
S/H
G
103
3
6
5
3
7
6
20
21
21
15
13
13
16
15
15
6
3
5
1.8
82
3
5
4
8
14
9
36
19
20
7
11
9
8
12
10
70
2
4
5
5
10
7
25
29
20
6
12
9
7
3
10
3
1
5
117
3
8
6
5
10
7
19
20
19
7
12
11
8
13
13
3
1
4
128
2
9
6
4
8
6
13
20
18
6
12
11
7
13
13
7
75
1
5
6
2
6
5
11
39
20
6
12
10
6
13
10
4
0.7
4
1.4
0.1
0.6
0.6
0.8
0.1
0.6
0.6
0.8
0.4
1.6
0.8
0.8
0.4
1.6
0.4
0.8
0.4
1.4
0.6
0.8
0.4
1.6
0.6
3
Note: F – Flaming phase; S – Smoldering phase; G – Global fire; B – Backing fire; H – Heading fire
In the case of CO2, the emission factors already integrate burning efficiency. Also, emission factors expressed as
burned fuel area for Mediterranean forest are presented in Table II.
Table II. Emission factors expressed as burned fuel area for Mediterranean forest (Simpson, 1999).
Emission factors (kg.ha-1)
Forest type
Mediterranean
forest
CO
CH4
NMVOC
NOx
NH3
N 2O
SO2
1456
54
133
51
11
3
11
Emission rates can be estimated using emission factors (g.kg-1) according to Ward and Radke (1993), which
depend on fuel load and consumption.
Emission ratios are other method to estimate forest fire emissions, specially used in field experiments, because
they do not depend of the burned mass fuel or composition. Emission ratio for pollutant Xi (RE[Xi]) is defined as
the ratio of the emission of pollutant Xi and the emission of other pollutant, chosen as reference. Usually, it is
chosen CO2 for flaming phase and CO or TP for smoldering phase, as reference pollutants.
Concentration ratios related to concentrations of CO, CH4 and PM2.5 are also used for emissions estimation for
forest fires, especially in assessment of human health exposure.
In the European context and according to data from the European emission inventory EMEP/CORINAIR
(European Environmental Agency, 2004), forest fire emissions represent 0.2% of nitrogen dioxide (NO2), 0.5%
of NMHC, 1.9% of CO and 0.1% of NH3. For Portugal the contribution of forest fire emissions in 2003 to the
total value equates to 14.1% CO, 5.2% NO2, 2.7% NMHC, 2.2% CH4, 1.3% NH3 and 0.6% SO2. These
emissions were estimated using the national emission inventory for non-forest-fire emissions and the model
36
EMISPREAD (Miranda et al., 2005b), which takes into account the type of fuel and combustion phase, to
estimate forest fire emissions from southern Europe.
2.2 Emissions measurement
Atmospheric emissions from forest fires can also be measured. The burning experiments performed since 1998
in Gestosa, Central Portugal, aim to collect a large range of different but complementary experimental data,
which can be used to support the development of new concepts and models and to validate existing methods or
models in various fields of fire management (Viegas et al. 2002), providing a particularly important opportunity
to measure and analyse air pollutant concentrations during experimental field fires (Miranda and Borrego 2002).
Experiments undertaken in 2004 are reported here in what concerns emissions measurement.
Borrego et al. (2005) describe the field experiments that have been carried out, the 2004 Gestosa experiments,
aiming to measure the atmospheric emissions due to vegetation burning. The vegetation was mainly composed
by shrubs with a relatively small height. The experimental area was divided into 15 plots with regular shapes and
dimensions (25 x 50 m2). Figure 1 presents the location of Gestosa 2004 and the burning plots, and Figure 2
shows two images of Gestosa 2004 fire experiments.
.
Figure 1. Map and schematic view of Gestosa 2004 plots and location of mobile laboratories, in Trevim
area.
Figure 2. Images from Gestosa 2004 fire experiments.
The measurement of smoke emissions was performed in order to evaluate the contribution of different fire stages
and burning conditions to the total Volatile Organic Compounds (VOC) emissions. The applied technique for the
measurement of VOC emissions consists on sampling the smoke into Tedlar bags using an appropriate pumping
device, as shown in Figure 3. Samples are then kept in a cool and dark environment. Subsequent laboratorial
analyses consist in submitting the samples to a gas chromatography technique with a flame ionisation detector
(FID).
37
Figure 3. Smoke sampling technique during smouldering phase (left) and example of a filled Tedlar bag
(right).
This technique allows sampling smoke during different fire stages (flaming and smouldering) and for vegetation
treated or untreated with retardant, enabling to determine specific emissions for different burning conditions.
Figure 4 shows two examples of the sampling procedure with and without retardant.
fire
progression
Figure 4. Central image represents a typical layout of the plots. Red area is the retardant strip and the
green one the vegetation. Left image shows the smoke sampling with untreated vegetation and right image
with treated one.
In Figure 55 the concentration of VOC for each plot, with and without retardant, is represented. The burns of
plots 712, 713 and 714 present a considerable increase on the amount of VOC emitted during the combustion of
the retardant treated vegetation.
VOC (mg.Nm-3)
6000
5000
4000
3000
2000
1000
0
700 701 702 703 704 705 706 707 708 709 710 711 712 713 714 715 716
Plot number
with retardant
without retardant
Figure 5. Representation of the concentration of VOC emitted for each plot with and without the presence
of retardant.
38
Almost all the samples registered very high values. The VOC peak measured concentration reached 1930 mg.m3
. The reason why the measurements were so high, particularly when compared to previous experiments
(Miranda and Borrego, 2002), is the fact that the sampling location was much closer to the burning vegetation
causing that the sampled smoke had lower dilution. Due to the small size of the plots and the low fuel load, the
intensity of the fire in 2004 was not as high as in previous years, allowing that the sampling could be made much
closer to the combustion.
3 Forest Fires and Air Quality
The above overview of the contribution of forest fires to total atmospheric emissions makes clear that some
relationship between forest fires and air quality is to be expected. However, it is not always easy to identify air
pollution episodes caused or exacerbated by forest fires. Pollutants emitted from forest fires are transported and
dispersed in the atmosphere and their effects on air quality can occur far from the emitting source. Although
major wildfires are limited to some hundreds of hectares, their impacts, with no natural or political boundaries,
can be felt and reported far beyond the physical limits of the fire spread. Depending on meteorological
conditions, smoke plumes and haze layers can persist in the atmosphere for long periods of time and prevailing
conditions will influence the chemical and optical characteristics of the plume.
The link between forest fires and air quality is not commonly made. From the point of view of the community
dealing with the fire the main concern lays in the direct effects of fire including human fatalities and property
damage. In addition, air quality problems are usually analyzed in terms of the main anthropogenic sources,
particularly the classic industrial and road transport sectors.
Evaluation of the effects of forest fires on the air quality can also be based on measurements or/and modelling
estimates.
3.1 Air Quality Measurements
During the experimental Gestosa 2004 fires two mobile laboratories (LM) were parked near the burning plots
(see Figure 1) and continuously measured PM2.5, PM10, NO2, nitrogen monoxide (NO), and CO levels. Figure 6
shows one of the mobile laboratories.
Figure 6. Mobile laboratory used during Gestosa experiments for air quality and meteorological data
acquisition. Outside (left) and inside (right) views.
Distinct techniques and equipment were used to obtain the concentrations of the different pollutants. The
description of these techniques and equipment can be found in Valente et al. (2007). In order to better understand
the experimental field fires’ effects on the air quality, the measured results were compared to European air
quality legislation values, which are also the Portuguese standards (see Table III).
Table III – Air quality limit values for the protection of human health established by European legislation.
Pollutant
Limit value
Averaging period
Directive of the Council
PM10
50 µg.m-3
24 hours
1999/30/EC
NO2
200 µg.m-3
1 hour
1999/30/EC
SO2
350 µg.m-3
1 hour
1999/30/EC
CO
10 mg.m-3
Maximum daily 8-hour mean
2000/69/EC
Figure 7 depicts the measured concentrations at LM1 of NO and NO2 (5 min average), and PM10 (15 min
average), respectively, for the second day of the experiments, when plots 702, 701 and 707 were burnt according
to the following schedule: 702 – ignition time: 09:50; 701 – ignition time: 10:10; 707 – ignition time: 10:50.
39
NO & NO2
Conc (ug.m-3)
700
600
500
400
300
200
100
0
8:24
9:36
10:48
12:00
NO
13:12
14:24
15:36
NO2
500
Conc (ug.m-3)
400
300
200
100
0
08:24
09:36
10:48
12:00
13:12
14:24
15:36
Figure 7. Measured concentration values of NO and NO2, and PM10 in LM1 on 12 May 2004.
The NO2 concentrations attained a significantly high peak of 275 µg.m-3. Nevertheless, the hourly averages
never exceeded the hourly European limit for NO2. When the plume reached the mobile laboratory, due to the
proximity with the emission location and its position downwind of the plume, the NO values were quite high,
reaching the 600 µg.m-3.
The calculated daily mean of PM10 concentration, 33 µg.m-3, did not surpass the limit value established by the
European legislation and only during approximately three hours this 50 µg.m-3 limit value was exceeded.
However, this direct comparison of measurements and European air quality thresholds has to be carefully
analysed, because only the monitoring effect of a small burning plot during a small period of time was taken,
while standards are established for longer averaging periods.
Although the burns only took place over a short period of time, the registered concentration values indicate that
levels of some concern are attained.
3.2 Air Quality Modelling
The extreme fire events occurred in the summer of 2003 in southern Europe highlighted the need to better
analyze the link between forest fires and air quality. In Portugal, the summer of 2003 was considered the most
devastating of the last decade in terms of forest fires, and this is clearly reflected in the values measured by the
air quality-monitoring networks (Miranda et al., 2005c). Several air quality stations registered extremely high
pollutant concentrations due to fire emissions and transport from surrounding areas.
Lisbon suffered the effects of smoke from forest fires north of its urban area in September 2003, particularly in
the 13th of September. The Lisbon airshed, with a population of 3.5 million inhabitants, is the most important
urban centre in Portugal. It was built in a very complex topographic region, dominated by a large estuary and
multiple hills and surrounded by small mountain ranges with elevations over 400 m above sea level. Because of
its urban/wildland characteristics, high population density, and hence higher risk of human exposure to smoke,
and the high levels of pollutants registered, Lisbon forest fires are a very interesting case for the study of the
influence of forest fires emissions on air quality.
Numerical modelling of smoke dispersion allows the understanding of how pollutants emitted by a forest fire are
transported and dispersed in the atmosphere by estimating the resulting air pollutant concentration fields.
40
AIRFIRE (Miranda, 2004) was developed to take into account the possible impact of forest fires on
photochemical production. Using data available for the 13th of September 2003, AIRFIRE was applied to a
modelling domain of 200 x 200 km with a horizontal resolution of 4 x 4 km (Figure 8). This domain was chosen
in order to consider mesoscale circulations, such as sea breezes in the Lisbon area. More details about this
particular case study can be found in Miranda et al. (2005c).
Altitude (m)
650
600
Pego
550
500
Montejunto
450
400
350
Sintra
300
Lisboa
250
200
Sines
150
100
Oceano Atlântico
50
0
Figure 8. AIRFIRE simulation domain.
Figure 9 shows surface PM10 hourly concentration values for two distinct situations: i) considering Lisbon
emissions for a normal week day; and ii) considering Lisbon emissions and emissions from the forest fires.
4m.s-1
0
30
60
80
100
200
300
400
500
600
800
1000
PM10
(µ g.m-3)
1250
1500
0
30
60
22:00
80
100
200
300
400
500
600
800
1000
1250
18:00
1500
i)
PM10
(µ g.m-3)
4m.s-1
200000
200000
180000
180000
Santarem
Santarem
160000
160000
140000
140000
120000
120000
Lisboa
100000
Setúbal
80000
Oc
ea
n
60000
40000
Lisboa
100000
Évora
Setúbal
80000
Oc
ea
n
60000
o
At
la
nt
ic
o
40000
o
Évora
At
la
nt
ic
o
20000
20000
0
0
0
20000
40000
60000
80000
100000 120000 140000 160000 180000 200000
0
20000
40000
60000
80000
100000 120000 140000 160000 180000 200000
41
0
30
60
80
100
200
300
400
500
600
800
1000
PM10
1250
1500
0
30
60
22:00
80
100
200
300
400
500
600
800
1000
1250
18:00
1500
i)
4m.s-1
(µ g.m-3)
PM10
4m.s-1
(µ g.m-3)
200000
200000
180000
180000
Santarem
Santarem
160000
160000
140000
140000
120000
120000
Lisboa
100000
Évora
Setúbal
80000
Oc
ea
n
60000
40000
Lisboa
100000
Oc
ea
n
60000
o
At
la
nt
ic
o
Évora
Setúbal
80000
o
40000
At
la
nt
ic
o
20000
20000
0
0
0
20000
40000
60000
80000
0
100000 120000 140000 160000 180000 200000
20000
40000
60000
80000
100000 120000 140000 160000 180000 200000
µg.m-3) concentration fields at 18:00 and 22:00 Local Standard Time
Figure 9 – Wind and PM10 (µ
considering i) only Lisbon emissions, and ii) Lisbon and fire emissions.
The influence from the fires on PM10 concentration is clearly visible, with larger pollutant clouds and
considerable higher concentrations.
The urban area of Lisbon has an air quality monitoring network that includes several stations with different
typologies (urban background or urban traffic), according to location and environmental criteria. Data from those
monitoring stations allowed evaluating the model performance when simulating the Lisbon air quality affected
by smoke from forest fires. Figure 10 presents the comparison between simulated and observed PM10 hourly
concentrations, for some of Lisbon’s air quality stations, for the 13th of September 2003.
Entrecampos
Av. Liberdade
SIMULATED
500
PM10 ( g.m-3)
PM10 ( g.m-3)
600
OBSERVED
400
300
200
100
0
1
3
5
7
9
11
13
15
17
19
21
450
400
350
300
250
200
150
100
50
0
23
SIMULATED
OBSERVED
1
3
5
7
9
TIME
13
15
17
19
21
23
15
17
19
21
23
TIME
Laranjeiro
Loures
400
250
200
SIMULATED
350
SIMULATED
OBSERVED
300
250
OBSERVED
PM10 ( g.m-3)
PM10 ( g.m-3)
11
150
100
50
200
150
100
50
0
0
1
3
5
7
9
11
13
TIME
15
17
19
21
23
1
3
5
7
9
11
13
TIME
Figure 10. PM10 (µ
µg.m-3) simulated and observed hourly concentration evolution in Av. da Liberdade,
Entrecampos, Cascais, Loures and Laranjeiro air quality monitoring stations.
A reasonable agreement between simulated and observed values was found. However, in Av. da Liberdade and
Laranjeiro stations simulated values were bellow measured ones, probably due to an under-estimation of fire
emissions associated to uncertainties in the factors involved in their calculation, namely fuel load and emission
factors.
42
4 Conclusions
The effect of forest fires on air quality is an issue of concern in many regions of the world, including the
southern European countries. Emissions from forest fires may cause substantial exceedances of the air quality
threshold and there is a strong need to take into account the role of forest fires when determining management
strategies for air quality. This manuscript illustrates how forest fires and air quality issues can be linked
describing two main approaches: measuring and modelling.
Experimental field fires represent a valuable tool for understanding wildfires in their all extension: behaviour,
impacts on environment, security conditions and health, suppression techniques efficiency, etc. The participation
in Gestosa fire experiments, particularly in the year 2004, has been a profitable opportunity to collect air
pollutant concentrations data. Distinct techniques and equipment were used to obtain the concentrations of the
different pollutants. Despite the small size of the burning plots when compared to real wildfires, the measured
levels of pollutants are not negligible. However, and this is one of the problems inherent to field experiments, the
exact conditions of each burn are not reproducible, in particular the meteorological conditions, the terrain
characteristics, the type and load of fuel.
The application of numerical air quality modelling systems is also an added value when evaluating and assessing
air quality levels in areas affected by forest fires. Fortunately, European researchers and managers are realizing
the importance of forest fires to the degradation of the air quality and some modelling works were already
developed and applied, e.g. Hodzic et al. (2007) and Miranda et al. (2007). These studies are simulating larger
areas, Europe and a European country, respectively, and longer periods (at least one month) than the one
presented here. This different modelling scale implies the simulation of a larger number of fires and the use of a
coarser grid resolution. It is a top down approach instead of the used bottom up approach of the presented case
study. Both modelling approaches can be useful. The bottom up can be applied to simulate a specific air
pollution episode related to the occurrence of forest fires. In this case the information from the numerical system,
which can be applied almost in real time, can help to identify critical areas where persons are exposed to high
levels of air pollutants. The top down approach is quite useful for the characterization and the evaluation of the
air quality by each European member state or at a higher level by the European Commission.
Finally, and as a last recommendation, to protect people, environmental policies must integrate the traditional
pollution-oriented and forested-land management issues into a unique system that can integrate both problems.
Better forested-land management can help to reduce both the number of air pollution episodes and the risk of
unwanted fires.
Acknowledgments
Part of this work has been developed within the Research Group “Emissions, Modelling and Climate Change”
from the University of Aveiro and a special thanks to Jorge Amorim, Joana Valente, Pedro Santos and Helena
Martins is due. The authors wish to recognize the importance of D. X. Viegas coordinating the experimental
Gestosa field work. Also, an acknowledgment is due to all the colleagues that participated in this work, and in
particular to IDAD for its technical assistance during the field measurements.
References
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of retardants. Deliverable D25 of the EC Project ERAS (Extension Retardant Application System).
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Hodzic, A., Madronich, S., Bohn, B., Massie, S., Menut, L. , Wiedinmyer, C. 2007. Wildfire particulate matter in
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Miranda, A.I., Borrego, C. 2002. Air quality measurements during prescribed fires. In ‘Proceedings of IV
International Conference on Forest Fire Research’, Luso-Coimbra, Portugal, 18-23 November, Forest
Fire Research & Wildland Fire Safety. (Ed. DX Viegas), Millpress, Rotterdam, Netherlands. pp. 205.
Miranda, A.I., Coutinho, M., Borrego, C. 1994a. Forest fire emissions in Portugal: A contribution to global
warming? Environmental Pollution 83(1,2), 121-123.
Miranda, A. I., Borrego, C., Viegas, D. 1994b. Forest fire effects on the air quality. In: Baldasano, J., Brebbia,
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vol. 1: Computer Simulation, Barcelona, Spain, pp. 191-199.
Miranda, A.I., Ferreira, J., Valente, J., Santos, P., Amorim, J.H., Borrego, C. 2005a. Smoke measurements
during Gestosa-2002 experimental field fires. International Journal of Wildland Fire 14, 107-116.
Miranda, A.I., Borrego, C., Sousa, M., Valente, J., Barbosa, P., Carvalho, A. 2005b. Model of forest fire
emissions to the atmosphere. Deliverable D252 of SPREAD Project (EVG1-CT-2001-00043). AMBQA-07/2005. University of Aveiro, Aveiro, Portugal.
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simulate smoke dispersion. In 6th Symposium on Fire and Forest Meteorology, Canmore, Alberta,
Canada, 25 - 27 October 2005 - Proceedings of 6th Symposium on Fire and Forest Meteorology.
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Merino, L., Miranda, A.I., Santos, P. 2002. Gestosa fire spread experiments. In ‘Proceedings of IV
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121
Ward, D. 1999. Smoke from wildland fires; in Health Guidelines for Vegetation Fire Events, Lima, Peru, 6-9
October 1998, Background papers, WHO.
Ward, D.; Radke, L. 1993. Emissions measurements from vegetation fires: a comparative evaluation of methods
and results; in Crutzen, P. and Goldammer, J. (eds.); Fire in the environment: The ecological,
atmospheric and climatic importance of vegetation fires. John Wiley & Sons, Chicester, England; pp
53-76.
Ward, D., Rothermel, R., Bushey, C. 1993. Particulate matter and trace gas emissions from the Canyon Creek
Fire of 1988. In: Society of American Foresters (Eds.), Proceedings of the 12th Fire and Forest
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WHO. 1999. Health guidelines for vegetation fire events. In: D. Schwela, J. Goldammer, L. Morawska and O.
Simpson (Eds.). World Health Organisation, Geneva.
45
46
MEDITERRANEAN FIRE REGIMES AND IMPACTS ON FOREST PERMANENCE
CASES FROM NORTHERN CALIFORNIA, NE SPAIN AND CANARY ISLANDS
Domingo MOLINA
University of Lleida - Department of Crop and Forest Sciences, Unit of Forest Fires
(UFF)
[email protected], [email protected]
Domingo Molina
MEDITERRANEAN FIRE REGIMES AND IMPACTS ON FOREST PERMANENCE CASES FROM
NORTHERN CALIFORNIA, NE SPAIN AND CANARY ISLANDS
DISCUSSING COSTS OF PRESCRIBED BURNING IN PRE-SUPPRESSION AND FOREST
MANAGEMENT. 6 YEARS OF EXPERIENCE IN CATALONIA (NE SPAIN)
Molina et al. 2007.
Since 1999, firefighters of “Generalitat de Catalunya”, NE Spain, are using prescribed burning in different
typologies of forest structures, and under different management goals which are within the scopes of preextinction and forest management. There have been 490 ha (table 1) treated by firefighters with prescribed fire as
a management tool since 1999. Additionally, we presume that this technique will continue to be used in the
future.
Pre-extinction congregates all treatment directed to transform or maintain a forest structure that can be used
safely and efficiently to anchor fire operations.
Forest management gathers different objectives; tree thinning, range improvement, habitat restoration for
wildlife and the fuel management (fuel breaks) to protect wildland interface areas.
Prescribed burning as a forestry tool has required a high effort in communication and training to overcome the
initial distrust and strangeness. In order to reach acceptance, a detailed assessment of costs and efficiency of both
traditional tools and fire should be carry out. The study is analyzing costs and productivity for different
objectives and structures treated. The final evaluation will be defined by the conditionings and restrictions
established which in turn depends on the management objective. This allows us to start a technical discussion
about productivity and profitability of prescribed burning to compare with other treatments.
Table 1. Number and area of prescribed burns by type that involved firefighters in Catalonia (NE Spain).
Prescribed burn type
number
number (%)
Area (ha)
Area (%)
Strategic sites for fire suppression
32
43
66
135
Fuel breaks
13
18
76
15.5
Industry Empty Lots / Urban Interface
5
7
6
1.3
Wildland Urban Interface
4
5
16
3.3
Research sites
2
3
2
0.4
Range improvement
16
22
324
66.2
Total
74
490
Mean burn size is 6.6 ha, ranging from research sites (< 1 ha) to range improvement sites where mean size is 20
ha
To assess both resources and safety needs, there was necessary to classify haw difficult a given prescribed burn
is. We consider this to be a combination of two factors a) fuel continuity (horizontal and vertical or fuel ladder)
and b) objective of the burn
The mean cost per hectare of prescribed burning is 1200 €/ha for major types (Strategic sites for fire
suppression, Fuel breaks and Range improvement that amount for 70% of all surface treated). Individual costs
and number of person ped day and hectare are displaied in Table 2. We can observed that range agement
objective reflect the best performance
Table 2. Cost and persons involved in each prescribed burn type that involved firefighters in Catalonia
(NE Spain).
Prescribed burn type
Cost (€/ha)
Person-day/ha
Strategic sites for fire suppression
1247
8
Fuel breaks
1379
10
Industry Empty Lots / Urban Interface
1148
9
Wildland Urban Interface
566
3
Research sites
2727
19
Range improvement
242
2
Keywords – prescribed fire, cost, pre-suppression, forest management
48
Mediterranean fire regimes and impacts on forest permanence cases from northern California, ne Spain and
Canary islands
COST-EFFICIENT WILDLAND FIRE PLANNING: CASE STUDIES IN SPAIN
Molina et al. 2007.
Wildland fire regimes have recently changed in Spain due to the growing weight of large wildland fires -LWF-.
Today, a much larger percentage of our area swept by fire belongs to the LWF fire class. We have more
suppression resources than before while LWFs are even bigger; therefore, our fire planning should change. Many
of our LWF are beyond the threshold of suppression forces. To fight these LWF, we have to build suppression
means (i.e., hydrants, pools, fuel break areas, dirt roads, firefighter safety zones) before suppression efforts. In
this study, we assess cost-efficient pre-suppression actions after careful study (multiple Farsite and Flammap
simulations) of the specific, most likely pattern of LWF behavior for each forest massif. In a like manner, we
locate the required aggressive wildland fuel management actions in priority areas (i.e., where suppression efforts
are useless with actual fuel conditions) using the above mentioned computer simulations. In this paper, we show
several case studies in both in NE Spain (Martinez 2002, Molina & Galan 1999, Molina et al 2004) and Insular
SW Spain. Firefighters needs are different in planning ahead (i.e., fuel management) and suppression actions
(Molina & Bardají 1998, Castellnou et al 2001, Castellnou et al 2002). A major need is to study and review
historical fires and to simulate their behavior (Finney 1998, Finney et al 1997, Molina & Castellnou 2000) to
learn to forecast fire behavior in future events. This should lead us to have a classification of LWF types to
properly assess the potential of future fires and have a proportional response to their potential. Therefore,
computer simulation tools are very useful for firefighters in planning ahead. On the contrary, in suppression
actions other tools are more helpful like Campbell Prediction System language (CPSL, Campbell 1995) to
advance most likely fire behavior as fire moves on the landscape and through time. This is especially true when
fires do not last many hours (most of the times). We show, among other cases, a study where treating (i.e., fuel
management) only 59.6 ha (strategically located) we are able to dramatically reduce the potential of large
wildland fire in 11309 ha of the study area in NE Spain. We have created a set of valuable opportunities to
contain (i.e., suppression fire o backfiring) a large wildland fire that respond to a LWF type that we have
previously characterize.]
A NETWORK OF PRESCRIBED FIRE DEMONSTRATION SITES FOR EUROPE, AFRICA AND
ARGENTINA
Molina, Goldammer, et al., 2007
Changes in wildland fire regimes and an increasing occurrence of large wildland fires are taking place or
expected for the near future. They constitute major changes to the current natural and anthropogenic disturbance
regimes and to biodiversity threats in several world ecosystems, thus reinforcing the potential mitigating role of
prescribed burning. Within the European Integrated Project FIRE PARADOX we are undertaking the ambitious
task of setting a network of prescribed fire demonstration sites covering ecosystems from subtropical Canary
Islands to Northern boreal Europe. We have also added to this network, some areas in Argentina and South
Africa. Included are pilot projects, experimental and demonstration sites for the application of prescribed fire in
the abatement of wildland fire hazard, forest succession management and ecosystem restoration We have agreed
on the methodology to choose, implement and run useful prescribed burning demonstration sites, both in forest
stands, non-forest vegetation and industrial plantations. Burn plots larger than 10 ha are a major priority, but
smaller burn sites would be considered due to limitations set by authorities in sensitive testing phases.
Experiences from the European Fire in Nature Conservation Network (EFNCN) and results from previous EU
projects like Fire Torch helped to gather a common minimum protocol for data collection and standardisation.
When available, images from the site before, during and after burning are used. In this paper, we show some
sites as case studies relating prescribed burns, changes in disturbance regimes and how prescribed burnt stands
have enhanced the efficiency of fire suppression operations.
•
KEYWORDS: prescribed fire, simulation, Farsite, FlamMap, pre-suppression
ASSESSMENT OF THE FOREST SERVICE PRESCRIBED FIRE PLAN: 5 YEARS OF EXPERIENCE
IN GRAN CANARIA, SPAIN
Fababú, Molina, et al 2007.
The science of forestry, as it pertains to forest fires, has undergone great changes in the last few decades.
Paradoxically, forest fires, which used to be considered an ecological catastrophe to be avoided at all cost, now
are considered to play a fundamental role in the dynamics of many ecological processes. In this framework, the
previous policies to suppress wild fires have shown ineffective in solving the problems of large fires caused by
great accumulation of forest fuel. These fires are characterized by being out of reach of any kind of suppression
system, due to an increased intensity and rate of spread, and danger to people, goods and ecosystems. Facing the
impossibility of eradicating the fires of our landscapes, the efforts of firefighters have been focused on avoiding
their most extreme behavior. Prescribed burning fits into this context as a tool of fuel management. Since the
year 2002 the “Servicio de Medio Ambiente del Cabildo de Gran Canaria” has been applying mechanical hand
tools, prescribed fire, and mix of both. This has been done not only with the objective of preventing fires, but
49
Domingo Molina
also for pasture maintenance and forest restoration. The total of the surface treated by prescribed burns was
166,6 ha in the period 2002-2006; this is, 0,27 % of the forest surface of the island. This practice puts Gran
Canaria in the forefront of the European territories in regards to forest fire management. This paper is a
retrospective of the last 5 years of experience in Gran Canaria, based on all previously gathered information so
far on prescribed burning carried out on the island, creating a database and also making a first analysis of their
main parameters.
PRESCRIBED FIRE USE TO ESTABLISH SHADED FUEL BREAKS: LAS MESAS DE ANA LÓPEZ,
VEGA DE SAN MATEO CASE STUDY, GRAN CANARIA, SPAIN
Molina, Fababú, et al 2007.
Wildland fire regimes are changing in Gran Canaria (Canary Islands, SW Spain) and large wildland fires (LWFs)
are more likely to occur. This is a major change in the disturbance regime and it is a real threat for biodiversity.
This paper studies prescribed burning as a mean to create fire resistant stands to allow wildland fire control. It is
based in the recent (2002-2005) prescribed burning program in Gran Canaria. We are interested not only in
setting fuel-breaks but in providing strategic locations for those (after Farsite and FlamMap simulations). In this
paper, we also show a case study for the Mesas (May 11th, 2002) wildland fire and the prescribed burn fuel
breaks established around it in the last years in which this fire was successfully anchor.
KEYWORDS: prescribed fires, simulation, Farsite, FlamMap, pre-suppression
PRESCRIBED FIRE AS AN EFFICIENT TOOL TO MOVE FROM ARTIFICIAL PINUS
CANARIENSIS PLANTATIONS TO CLOSE TO NATURAL FOREST STRUCTURES IN GRAN
CANARIA, SPAIN
OR
FIRE MANAGERS TRYING TO RESEMBLE NATURAL DISTURBANCE PROCESSES TO ALLOW
FOR FOREST SUSTAINABILITY
Arévalo, Molina, et al., 2005.
Wildfire in the Pinus canariensis forest of Tenerife and Gran Canaria are of ecological significance, but has been
little studied. Fire is considered an ecological catastrophe that should be prevented. The present study was
designed to report the effects of fire effects on
understory species composition as a means to evaluate this premise. Due to the effect of the site in the species
composition, we suggest the use of multivariate analysis of the species composition under the effect of fire. P.
canariensis were planted in the sites during the 40s and 50s for reforestation purposes, but no other management
activities have been carried out in the next 30 years. The pine forest is a fire-prone ecosystem within which
species have a high ability to regenerate after fire. Although the rate of fire has on the whole increased in recent
years, the affected area is on average much lower. Knowing this, our results suggest that a regular occurrence of
fire as an internal process of the ecosystem will favor and accelerate the change of the pine stands to more
natural forests, and this is the aim of the reforestation programs. Therefore, prescribed fire is a tool to help
foresters in this restoration process.
KW: Species composition, DCA, disturbance, succession, prescribed fire, restoration.
PRESCRIBED BURNING EFFECTS ON SOIL WATER INFILTRATION RATES IN NORTHERN
CALIFORNIA SOILS AT BOGGS MOUNTAINS STATE DEMONSTRATION FOREST
Molina 1993.
50
Canary islands
MORE ABOUT PRESCRIBED FIRE TECHNIQUES
This is based on USA standards (http://www.tncfire.org/admin.htm) and outcomes from a several EU research
projects i.e., BOTELHO et al, 2002 and Molina et al. 2007
Develop or Update the Prescribed Burn Plan and Checklist
What is a Prescribed Burn Plan and Checklist?
The Prescribed Burn Plan is the most significant document in the burning planning process. It is required for all
burns. It is a field document that sets forth the details for conducting a specific burn treatment at a particular burn
unit. The Prescribed Burn Plan is much more specific than the Site Fire Management Plan. It is also a legal text
which details the professional standards and guidelines to be used in conducting the burn. The Prescribed Burn
Plan includes:
•
•
•
•
•
•
•
•
the objectives to be accomplished by a particular prescribed fire
an satisfactory range of fire weather, fuel moisture, and fire behavior parameters to safely achieve
desired effects
burn-specific information on hazards, smoke susceptible areas, and contingencies, escape routes and
security zones
details of pre-burn site preparation, probable ignition patterns, crew assignments, holding positions, and
mop-up activities
lists of equipment needed for the burn
sources of emergency assistance
a series of high-quality maps showing the site/preserve, burn unit, smoke sensitive areas, ecologically
sensitive areas, proposed ignition pattern, and escape routes, safety zones and secondary control lines
a checklist for burn preparation and crew briefing
Many burns require a new or revised Prescribed Burn Plan, and every Fire Boss should complete a checklist
prior to igniting any approved burn. If there are no changes in the Burn Plan from one burn to the next, it is not
necessary to resubmit a new one for each subsequent burn once the Plan has been approved. If there is any
51
Domingo Molina
question about the need for formal review of an update, check with the Fire Manager. A prescribed fire is
authorized only if the Prescribed Burn Plan is approved and all conditions stated therein are in effect.
Who Prepares and Approves a Prescribed Burn Plan?
The Prescribed Burn Plan and Checklist is usually prepared by the Fire Boss who will be conducting the burn,
but may be prepared by a Fire Planner. Regardless, in-depth familiarity with the fire unit and material presented
in this manual is essential to successfully completing the Prescribed Burn Plan. All Plans must be signed and
dated by the preparer, by a Fire Manager, and by the Fire Leader. Usually, the person who authors a Prescribed
Burn Plan cannot be the sole person to approve it. In a limited number of situations, where a site/preserve has a
detailed Site Fire Management Plan and burns are conducted routinely by an experienced Fire Leader, the Fire
Manager may authorize the Fire Leader to sign off on the Prescribed Burn Plan without further review and
approval.
Is There a Standard Prescribed Burn Plan Format?
The standard Prescribed Burn Plan format is available for download on the National Fire Management Program
Internet web page (http://www.tncfire.org/admin.htm). You may customize the form as long as it meets with the
approval of your Fire Manager. All fields in the prescription form must be filled. Where the information is not
applicable to your situation, this should be stated. Already existing Prescribed Burn Plans using formats other
than this stand approved and do not have to be recast until major revisions are necessary..
Contractors or public agency cooperators working for the Conservancy may have versions of Prescribed Burn
Plans covering essentially the same information. Different formats may qualify for this planning requirement
provided they include the same types of information as the standard Conservancy plan.
Can Prescribed Burn Plans Be Modified in the Field?
Yes, under certain limited instances the Fire Leader may make a decision in the field on the day of the burn to
modify the conditions or criteria stated in the Prescribed Burn Plan. There is a place on the Checklist to note
changes in the Burn Plan and to state a justification. All field modifications must be justified in writing and
signed by the Fire Leader. An example might be that a downwind field on adjacent property was recently plowed
creating a more secure base line that allows a wider prescription window. Under no circumstances should any
changes be made the day of the burn that do not follow the Prescribed Burn Guidelines.
Where are Prescribed Burn Plans Filed?
Copies of Prescribed Burn Plans should be archived at the state or preserve level and with your Fire Manager. A
copy is also carried into the field during the prescribed burn.
How Are Burn Objectives Met?
The process of meeting burn objectives involves identifying and applying the appropriate fire behavior to
produce the desired effect while considering the feasibility of obtaining and managing such a fire.
State burn objectives so that they can be objectively assessed. Burn objectives apply to a particular burn
treatment. Examples are: remove 70% or more of the surface litter; reduce shrubs from 80% canopy cover to a
range of 20 to 40%; remove duff to expose mineral soil over 15 to 30 percent of the ground surface to provide
suitable seed bed for certain species. The burn objectives should be specific enough for the planner to assess
what kind of fire is required.
Once the objectives are clearly stated, the next step is an initial approximation of prescribed fire behavior. The
problem is to go from an objective such as "remove 70% or more of the surface litter", to fire characteristics
required to accomplish that objective. "A fire that will cover 70% or more of the surface area and consume
virtually all surface litter" is a start at this approximation. [Tools to help the planner are described here. Also see
references here.] These tools can help answer questions such as: how dry must fuels be to burn readily and what
flame lengths and rate of spread would be associated with a fire that would consume them given certain fuel
moisture conditions? Prediction tools, however, are not a substitute for experience. In general, the best way to
develop an approximation of what fire characteristics are needed to accomplish burn objectives is to talk to
52
Canary islands
people experienced with fire effects and behavior in the community type to be burned. Published literature may
also be helpful.
Next, consider the feasibility of producing the approximated fire behavior and managing a fire of this nature.
Begin with an assessment of fuels on the site, and likely weather conditions during the season targeted for
burning. Then estimate fire behavior under projected weather and fuel conditions using fire behavior prediction
models and guides or other methods.
The process is complete when the Prescribed Burn Plan is refined to a point where your objectives can be met
given the probable fuel and weather conditions, yet the fire can be safely conducted given legal, personnel, and
equipment constraints. The result is a Prescribed Burn Plan that
•
•
•
•
•
includes fire behavior ranges for flame length, rate of spread, and scorch height
prescribes ranges for temperature, relative humidity, wind speed, wind direction, atmospheric stability
and fuel moisture of different classes and sizes of fuel
defines personnel and equipment needs
clearly outlines safety and contingency planning
describes ignition, holding and mop-up activities
AFTER THE BURN
Monitor prescribed burning effects
Monitor fire effects and species and community responses as outlined in your monitoring plan. If you need
assistance, contact the Conservancy's National Monitoring and Adaptive Management Program
Refine Plans and Burn Again
It is important at this stage to use the fire summary report and monitoring data to assess progress toward meeting
the site fire management goals. You should then refine the Site Fire Management Plan and/or Prescribed Burn
Plan if necessary before entering another cycle of burning and monitoring.
53
Domingo Molina
References
Arévalo, J.R., Molina-Terrén, D., Naranjo, A., García-Marco, D., Grillo, F., Velázquez-Padrón, C., 2005.
Naturalización de las masas de Pinus canariensis mediante fuego prescrito. Libro de Resúmenes del
Congreso Forestal Español. Zaragoza.
Botelho, H., Fernandes, P., Rigolot, E. Rego, F., Guarnieri, F., Binggeli, F., Vega, Ja., Prodon, R., Molina, D.,
Gouma, V., Leone, V. , 2002. Main outcomes of the Fire Torch Project : a management approach to
prescribed burning in mediterranean Europe. 4. International Conference on forest fire research / 2002
Wildland fire safety summit, Luso (PRT), 2002/11/18-23; PIF 2002-13, feu=5243 http://www.eufirelab.org/toolbox2/library/upload/224.pdf
Cuiña, P. 2002., Manejo del fuego como herramienta selvícola. Posibilidades en las masas de Pino Canario. In
Actas de las Jornadas forestales de Gran Canaria 1994 - 2001. Cabildo de Gran Canaria. Las Palmas de
GC.
Fababú, D.D., Grillo, F., García-Marco, D. Y, Molina-Terrén, D.M. 2007. “Caracterización de las quemas
prescritas en Gran Canaria. Valoración de 5 años de experiencia” Revista incendios forestales. Abril
2007:4-17 http://www.incendiosforestales.com/if16.pdf
Finney, M.A., 1998. FARSITE: Fire Area Simulator - model development and evaluation. USDA For. Serv. Res.
Pap. RMRS-RP-4. 47p
Finney, M.A., 1999. FlamMap. USDA For. Serv. Rocky Mountain Research - Missoula Fire Lab.
García-Marco, D., Grillo, F., 2003. Plan de Prevención de Incendios de Gran Canaria. Cabildo de Gran Canaria.
Las Palmas de GC.
Larrañaga, A., Galán, M., Pellisa, O., 2005. Discusión sobre el análisis de costos de las quemas prescritas en los
ámbitos de pre-extinción y gestión forestal. Valoración de 6 años de experiencia en Cataluña. II
Conferencia internacional sobre estrategias de prevención de incendios en el sur de Europa. Barcelona.
Martínez, E., Larrañaga, A., 2004. Programa de gestió de cremes prescrites a Catalunya. In: Plana E. (Ed.)
Incendis forestals, dimensió socioambiental, gestió del risc i ecologia del foc. Zarza ALINFO
XCT2001-00061. Solsona, DL: L-501/2004, 144 pp.
Molina, D.M. 1993. Efectos del fuego controlado en la velocidad de infiltración del agua en suelos forestales: un
caso de estudio en la costa norte de California. Revista de Investigación Agraria: Sistemas y Recursos
Forestales, 2(2):173-84. INIA, Madrid.
Molina-Terren, D.M., Martinez-Lopez, E.R., Garcia-Marco, D. 2006. Farsite simulations for cost-efficient
wildland fire planning: Case studies in Spain. Proceedings of V International Conference on Forest Fire
Research held in Figueira de Foz (Portugal). D. X. Viegas (Ed.). Forest Ecology and Management 234S
(2006) S217.
Molina, D.M. , Galán, M., Fababú D.D., García, D., Mora, J.B. 2007. Prescribed fire use for cost effective fuel
management in Spain. IV International Wildland Fire Conference, Seville, Spain http://www.eufirelab.org/toolbox2/library/upload/2206.pdf
Molina, D.M., Grillo-Delgado, F., Garcia-Marco, D., 2006. Uso del fuego prescrito para la creación de rodales
cortafuegos: estudio del caso “Las Mesas de Ana López”, Vega de San Mateo, Gran Canaria, España.
Invest Agrar: Sist Recur For (2006) 15(3), 271-276, Madrid
http://www.inia.es/gcontrec/pub/271-276-(15)-uso_del_fuego_1166008640062.pdf
Molina, D.M., Fababú, D.D., Grillo, F.F., García, D., Arévalo, J.R. 2007. Uso del fuego prescrito para establecer
cortafuegos sombreados: caso de estudio de Gran Canaria - Prescribed fire use to establish shaded fuel
breaks: case studies in Gran Canaria, Spain. IV International Wildland Fire Conference, Seville, Spain http://www.eufirelab.org/toolbox2/library/upload/2226.pdf
Molina-Terrén, D.M., Grillo, F., Fababú, D.D., 2006. Mesas (May 11th, 2002) wildland fire (Spain) of and the
effect of the wise use of fire (both prescribed burning and suppression fires) in its control.
(http://www.fireparadox.org).
Molina-Terrén, D.M., Grillo, F., García-Marco, D. 2006. Uso del fuego prescrito para la creación de rodales
cortafuegos: estudio del caso “Las Mesas de Ana López”, Vega de San Mateo, Gran Canaria, España.
Invest Agrar: Sist Recur For (2006) 15(3), 271-276, Madrid.
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Canary islands
Molina-Terrén, D.M., Grillo, F., García-Marco, D., Fababú, D.D., Velázquez-Padrón, C., 2005. Establecimiento
de rodales resistentes al paso del fuego con el empleo del fuego prescrito. Libro de Resúmenes del
Congreso Forestal Español. Zaragoza.
Molina-Terrén, D.M., Martínez-López, E.R., García-Marco, D., 2006. Farsite simulations for cost-efficient
wildland fire planning: case studies in Spain. V International Conference on Forest Fire Research, D. X.
Viegas (Ed.)
Molina, D.M. 2000. El fuego prescrito. En: Incendios Forestales: Fundamentos y Aplicaciones (Vélez, R. ed.).
McGraw-Hill, p. 14.36-9.41. ISBN 84-481-2742-0
Molina, D.M., Goldammer, J.G., Loureiro, C., Castellnou, M., Vega, J.A., Delogu, G., Rigolot, E., Defossé, G.,
Kunst, C., de Ronde, C., Abdelmoula, K., Sesbou, A. 2007. A network of prescribed fire demonstration
sites for Europe, Africa and Argentina. IV International Wildland Fire Conference, Seville, Spain http://www.eufirelab.org/toolbox2/library/upload/2168.pdf
Molina, D.M. 2000. Planes de quemas. Prescripciones. En: Incendios Forestales: Fundamentos y Aplicaciones
(Vélez, R. ed.). McGraw-Hill, p. 14.42-14.60. ISBN 84-481-2742-0
Rothermel, R.C. 1983. How to predict the spread and intensity of forest and range fires. USDA Forest Service,
General Thecnical report INT-143. Ogden, UT, USA.
USA National Fire Management Program, a Prescribed Burn Plan format (http://www.tncfire.org/admin.htm)
55
56
REMOTE SENSING FOR POST FIRE FOREST MANAGEMENT
Gherardo CHIRICI
Dipartimento di Scienze e Tecnologie per l’Ambiente e il Territorio, Università del Molise
[email protected]
Gherardo Chirici
REMOTE SENSING FOR POST FIRE FOREST MANAGEMENT
Abstract
Monitoring of forest burnt areas has several aims: to locate and estimate the extent of such areas; to assess the
damages suffered by the forest stands; to check the ability of the ecosystem to naturally recover after the fire; to
support the planning of reclamation interventions; to assess the dynamics (pattern and speed) of the natural
recovery; to check the outcome of any eventual restoration intervention. Remote sensing is an important source
of information to support all such tasks. In the last decades, the effectiveness of remotely sensed imagery is
increasing due to the advancement of tools and techniques, and to the lowering of the costs, in relative terms. For
an effective support to post-fire management (burnt scar perimeter mapping, damage severity assessment, postfire vegetation monitoring), a mapping scale of at least 1:10000-1:20000 is required: hence, the selection of
remotely sensed data is restricted to aerial imagery and to satellite imagery characterized by high (HR) and,
above all, very high (VHR) spatial resolution. In the last decade, HR and VHR passive (optical) remote sensing
has widespread, providing affordable multitemporal and multispectral pictures of the considered phenomena, at
different scales (spatial, temporal and spectral resolutions) with reference to the monitoring needs. In the light of
such a potential, the integration of GPS field survey and imagery by light aerial vectors or VHR satellite is
currently sought as a viable option for the post-fire monitoring.
1. Introduction
Designing post-fire management (reclamation and restoration interventions, etc.) in forest burnt stands requires
the functional and structural assessment of the landscape mosaic and of the size and behaviour of the considered
wildfires. In this framework, monitoring may have several aims to support post-fire management: to locate and
estimate the extent of the burnt areas (burnt scar perimeter mapping e.g., Eva & Lambin 1998, Smith et al. 2002,
Holden et al. 2005); to assess the damages suffered by the forest stands (e.g., McHugh & Kolb 2003); to check
the ability of the ecosystem to naturally recover after the fire (e.g., Henry & Hope 1998, Diaz-Delgado et al.
2003, Moya et al. 2007); to support the planning of reclamation interventions; to assess the dynamics (pattern
and speed) of the natural recovery (e.g., Turner et al. 1997, Davis et al. 2005); to check the outcome of any
eventual restoration intervention (e.g., Raftoyannis & Spanos, 2005).
Remote sensing is a relevant source of information to support all such tasks. However, it must be taken into
account that in most Mediterranean countries the average size of forest fires (1980-2005) is currently less than 10
ha. In European central and northern countries the average burnt area is even much smaller, generally less than 1
ha (EC 2006).
During the last summer (year 2007) in the Mediterranean area, especially in Greece and in Italy, a lot of wildfires
occurred with a significant impact because of the wide range of economic, political, social and ecological values
at stake. At the end of August 2007, Greece asked to activate the International Charter “Space and Major
Disasters”, where national space agencies and private satellite companies make available remotely sensed data to
quickly manage human or natural disasters. In the light of this, satellite images with low (e.g., Noaa Avhrr,
Modis, Spot-Vegetation) or medium (e.g., WiFS-IRS) spatial resolutions may be exploited for monitoring the
largest wildfires, as, for instance, it is carried out by the European Forest Fire Information System (EFFIS, see
EC 2006) whit reference to those ones larger than 50 ha using Modis satellite images.
As concerns post-fire forest management, a mapping scale of at least 1:10000-1:20000 is required for an
effective support to operational measures like legal constraints enforcing and reclamation and restoration
interventions: hence, the selection of remotely sensed data is restricted to aerial imagery and satellite imagery
characterized by high (e.g., Spot HVR, IRS LissIII, Aster) and, above all, very high (aerial imagery, Ikonos,
QuickBird, Spot 5 and the upcoming WorldView-1 and GeoEye-1) spatial resolution.
For instance, in Italy the national regulation about forest fires (National Law no. 353/2000) requires the
implementation of a cadastral geodatabase of the burnt areas. For such a purpose, the boundary of all the burnt
forest stands have to be overlaid with digital cadastral maps in order to identify land ownership and to prevent
possible illegal post-fire activities such as land use changes (typically, for grazing or settlements). In this context,
the Italian National Forest Service has decided to map burnt forest stands at a scale of at least 1:5000 (minimum
mapping unit equal to 625 m2), in addition to the historical archives censued since 1970.
The aim of this note, which update and widen a previous overview (Chirici & Corona 2005), is to selectively
present, as a commentary discussion, relevant experiences in exploiting remotely sensed imagery to operatively
support post-fire management, with distinctive reference to burnt scar perimeter mapping and damage severity
assessment, by case studies mainly from Italy.
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Remote sensing for post fire forest management
2. Burnt area mapping
Detailed burnt area mapping (reference scales: 1:5000-1:25000) may be operatively based on: exclusive use of
field data (GPS, laser station, etc.), exclusive use of remotely sensed data, data integration (remotely sensed +
field data).
2.1. Field-based survey
Burnt scar perimeter mapping based on field work only (walking along the boundaries of the burnt area with a
GPS device) is theoretically the best option, being a rather objective way to directly acquire the field truth.
However, such an approach may have several shortcomings: possible technical problems from GPS e.g., because
of radio signal lacking due of satellite position; residual vegetation coverage; orography; rainfall; etc.;
difficulties in identifying burnt scar perimeter in the case of patchy fires; low accessibility for a safe work of the
field team; large extension of the burnt stands.
When such shortcomings becoame operatively relevant, then the following options for the field work may be
considered:
- field mapping based on digital orthophotos supported by a digital based topographic map (at least at 1:10000
scale) loaded on a palm or tablet PC connected with GPS: in a standard GIS environment the operator can
digitise manually the boundary of the fire visually localized;
- proximal sensing: the method is based on the availability of a multiple station made by a GPS and a laser
collimeter connected together by a palm or tablet PC; the operator, located on a favourable point of view, looks
at the burnt area, registers its position by GPS and then acquires the coordinates of some relevant points of the
burnt area with the laser instruments; such a procedure is particularly suitable in case of low accessibility on the
burnt stand, when fire boundary is rather regular or even in the case of patchy fires (in latter case, a distant point
of view can simplify the identification of the scar perimeter); however, the procedure requires specific
equipment and technical skill and it has to be carried out just after the event, especially under Mediterranean
conditions where vegetation regrowth is very fast.
2.2. Remote sensing-based survey
Burnt areas have a typical spectral signature, especially if analysed by a multitemporal approach, because of the
different ground coverage between pre-fire (vegetation) and post-fire (white ash, black ash, bare soil, dead
vegetation) conditions (Lentile et al. 2006).
Aerial photography has an high efficiency potential for detecting and delineating burnt areas, especially in the
case of light aerial vectors and in the case of sensors capturing infrared (IR) and near infrared (NIR) reflectances.
However, in the last decade the most growing segment of remote sensing is that of passive (optical) multispectral
imagery from satellites for earth observation (EO), due to the relative lowering of the costs and increasing spatial
resolution (Wuldner & Franklin 2003). Almost all the different types of available passive and active EO imagery
have been tested for experimental tests and/or for developing operative services for fire risk forecasting, near real
time alarm and support for fire fighters, burnt area mapping, quantification of burnt biomass, evaluation of fire
severity and damage assessment, and vegetation regrowth monitoring.
As mentioned, for most post-fire operational applications, high (HR) and, above all, very high (VHR) resolution
data are particularly suitable today: they are characterized by improved information depth (over 8 bit per pixel
per spectral band), fairly high signal-to-noise ratios and allow mapping with a consistent geometrical reliability
for management purposes (e.g., panchromatic channel: 10 m for Spot HRV; 6 m for IRS-1C and Spot 4; 2.5 m
for Spot 5; ≤1 m for Quickbird and Ikonos; ≤0.5 m for the upcoming WorldView-1 and GeoEye-1).
Image classification methods may be based on manual photointerpretation or on automatic/semiautomatic image
processing algorithms (Franklin 2001, Lillesand et al. 2004). The approach may be mono-temporal, based on the
acquisition of just a single post-fire image, or multi-temporal, based on the acquisition of at least two images,
one pre-fire and one post-fire.
Monotemporal approach is relatively cheap and of easy technical implementation. However, the identification of
burnt areas may be difficult if the image is acquired during the spring or it is temporally distant from the event.
In such situations, the vegetation regrowth may change the spectral signature of the burnt area which may
become similar to that of the surrounding unburnt land cover. Classification problems may also arise when forest
and non-forest fires have to be distinguished. The monotemporal approach is commonly applied by HR/VHR
data. Distinctively, object oriented-based techniques can significantly improve the detection of burnt area
mapping by VHR data (Gitas et al. 2004, Mitri & Gitas 2004, 2006). In the case of medium-low resolution data
(pixels of size between 180-250 m and 1 km), burnt areas pixels frequently have a mix signature with averaged
digital numbers corresponding also to unburnt ground coverages.
In the case of multitemporal approach, based on the acquisition of (at least) two images (pre- and post-fire), the
identification of burnt area is as easier as the post-fire image is acquired temporally closer to the event. The
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Gherardo Chirici
classification of the area based on the type of vegetation burnt is more feasible than the monotemporal approach.
However, the multitemporal method is more expensive and technically complex to be implemented, since all the
images have to be geometrically co-registered and spectrally normalised.
Multitemporal low and medium resolution EO data are frequently used for monitoring big forest fires over large
territories (e.g. Stroppiana et al. 2003, Csiszar et al. 2005). HR/VHR imagery is usually exploited to support
post-fire management by: (i) comparison of the digital numbers from selected bands or their combination
(vegetation indices), before and after the fire; (ii) comparison of the classifications independently carried out
before and after the fire.
Beyond the well known NDVI (Normalized Difference Vegetation Index) and IRI (Infrared Index), which
exploit the reflectances in the red and NIR bands, the most effective vegetation indices for forest burnt area
mapping prove to be NBR (Normalized Burned Ratio), based on red and SWIR bands, and NBRT, a
modification of the latter which includes also the so called “scaled brightness temperature” derived from the
thermic band (Holden et al. 2005).
The multitemporal quantitative comparison can be normalised through the MNI index, ranging between –1 and
+1, MNI = (imagepost-imagepre)/(imagepost+imagepre), where image is a vegetation index or the digital number of a
given band, imagepost is acquired after the fire event and imagepre is acquired before the fire event. MNI index
may be applied in order to identify burnt areas also on the basis of post-fire regrowth: acquiring imagepre just
after the fire and imagepost some time after, in the resulting MNI image the burnt areas are usually characterized
by significantly higher values than the surroundings (Chirici & Corona 2005).
2.3 Operative case studies
2.3.1 Cadastral geodatabases of burnt areas in Campania (Southern Italy)
Regione Campania has implemented a cadastral geodatabase of burnt areas for the years from 2000 to 2003 in
order to answer legal specifications settled by the already mentioned Italian regulation on forest fires (Law no.
353/2000).
The mapping procedure was based on a two-step approach. In a first phase, for each year an unsupervised
analysis of multitemporal Landsat 7 ETM+ imagery was carried out in order to identify those areas with a
spectral behaviour compatible with a forest fire event. For such a task, the MNI index based on band 4 (NIR
channel) of Landsat 7 ETM+ sensor was determined as the best tool for an unsupervised approach, confirming
previous experiences (Bottai et al., 2000). In the MNI image, burnt areas have lower values than unburnt ones
because of decreased level of photosynthetic activity (Figure 1). However, such a spectral behaviour is
compatible also with other events such as forest clearcuttings that, on the other hand, may be distinguished
because of their regular geometric shapes. For each year of analysis several MNI images were elaborated in
order to automatically identify burnt areas also on the basis of post-fire photosynthetic decrease and regrowth
photosynthetic increase. The final result is a reliable database of possible burnt areas.
In the second step, each possible burnt area were assessed also combining the geodatabase of field surveys
acquired by the National Forest Service. If a possible fire was confirmed by official field report, the perimeter of
burnt area was digitised manually at a reference scale of 1:10000 on the basis of digital orthophotos, yearly
acquired, and digital topographic maps. The resulting final geodatabase of burnt areas was finally overlayed with
digital cadastral maps in order to identify landowners for the application of the National Law no. 353/2000. A
total number of 19 Landsat images was elaborated in order to identify 2080 forest fires, for a total burnt area, in
four years, of 23068 ha (Grasso et al. 2004).
60
Figure 1: Example of the behaviour of MNI index to identify burnt forest stands in Campania (Central
Italy) by the near infrared band from Landsat 7 ETM+ images. Above, MNI index elaborated from a
post-fire image acquired the 2nd of August 2000 and a pre-fire image acquired the 1st of July 2000: the
areas with lower photosynthetic activity are darker. Below, MNI index for the same area from a post-fire
image acquired the 2nd of August 2000 and an image acquired during the following growing season (1st of
May 2001): the areas with increasing photosynthetic activity are lighter. White circles in the above image
identify the burnt areas.
2.3.2 European Forest Fire Rapid Damage Assessment System
At the European level, the EFFIS service (see § 1) analyses a multitemporal database of medium resolution EO
images to map forest fires larger than 50 ha for each Mediterranean country (Portugal, Spain, France, Italy,
Greece), every year. All the information is stored in a module referred to as EFFIS Rapid Damage Assessment
(RDA) and it performes the evaluation of damages at least twice during a fire campaign. The service is
implemented by the Joint Research Centre of the European Commission of Ispra (EC-JRC) since 1997
supporting the Forest Focus regulation (EC-2152/2003) (EC 2006).
The RDA procedure is based on the multitemporal analysis of MODIS images acquired during the vegetation
season in order to identify the events. In a second step, burnt areas boundaries are overlaid with Corine Land
Cover vector database to identify burnt land covers. For each country quantitative relationships have been
established between fire statistics from national authorities and large forest fires mapped by RDA module.
Hence, the EFFIS RDA system is able to estimate yearly the total forest burnt area on a national level for the
considered countries, with good accuracy (EC 2006). However, the produced figures have only a statistical
meaning, and even for large fires they cannot be used for burnt area monitoring at the detailed scale required by
post-fire management (legal framework, reclamation and restoration activities, etc.).
2.3.3 Object-based approach
Object-based techniques are particularly efficient to exploit the information by HR/VHR remotely sensed data
(Benz et al. 2004). They have been successfully used to accurately map burnt forest stands in many
Mediterranean areas. Mitri & Gitas (2006) developed an object-oriented model using a post-fire Ikonos imagery
of Thasos island (Greece) to distinguish between surface and canopy burn by a monotemporal approach: the
61
Gherardo Chirici
overall accuracy of classification was 0.87. A similar level of classification accuracy is reported by Chirici &
Corona (2006) with reference to events occurred in northern Latium (central Italy) assessed by MNI index
applied to multitemporal QuickBird images. Chirici et al. (2006) also report the remote sensing assessment of
post-fire forest recovery after a large wildfire occurred in the Cilento National Park (Southern Italy); the
recovery dynamics was monitored by various approaches applied to multitemporal Landsat images from 1993 to
2004: the most accurate results were obtained by multiresolution segmentation and object-oriented classification,
compared to pixel-based approaches. De Matteo et al. (2007) reported the possible operative use of segmentation
techniques applied to QuickBird post-fire imagery in Alpine areas, with errors ranging from 3 to 20% in respect
to field assessment.
Figure 2 shows an example of multiresolution segmentation of an Ikonos image of a Mediterranean forest stand
six years after the fire. The obtained segments (polygons) delineating different patches of recovering vegetation
are of direct support for stratification restoration planning (no intervention, seeding, planting, etc.).
Figure 2: Multiresolution segmentation of an Ikonos image of burnt forest stands after six years from the
event (Castel Fusano, Rome, Italy).
3. Damage assessment
Multispectral remotely sensed images are of great potential for classifying and mapping the severity of fire
damages to forest vegetation. The overall view of the burnt areas and the possibility of analyzing multispectral
reflectance make such imagery able to detect and qualify damages even better than visual assessment by field
teams (Franklin 2001, Wang et al. 2004).
62
3.1. Operative case studies
3.2.1. Aerial assessment: the SIMiB project
The SIMiB project (System for Mapping Forest Fires), funded by the Italian National Forest Service, has
tested the Duncan MS4100 (Duncan-Flir-Riegl) sensor, a multispectral tool carried by an aerial light
vector (SKY-ARROW 650TC). The sensor is based on three CCD cameras active in the visible and nearinfrared bands and one active in the thermal infrared band. Infrared false colour digital images are
produced and orthocorrected with a nominal accuracy of +/- 1 m in order to identify burnt area boundary,
to map damage levels and to estimates emissions of CO 2 caused by forest fire and hypotethical recoils on
Kyoto Protocol (Carlini et al. 2006). An example of a similar device suitably carried by light aerial vector
for post-fire assessment is the ASPIS sensor (Figure 3) developed by the University of Tuscia (Italy). The
methodology has proven to provide successfully monitoring results over 2000 km2 in Central Italy (Lazio)
allowing to map forest burnt areas with an accuracy compatible with 1:10000 scale level. Remote sensing by
aerial light vectors can be efficiently (in terms of imagery quality and cost) used to map and qualify with high
accuracy damages on distinctive spots, and it is also particularly suitable for calibrating remotely sensed imagery
of lower geometric resolution (Oertel et al. 2003).
Figure 3: Light aerial vector Sky Arrow 650 TC and the ASPIS sensor.
3.2.2. Damage assessment by VHR satellite imagery
Data acquired by the first commercial VHR satellite (Ikonos) have been exploited in the framework of post-fire
forest management activities carried out in the woodland of Castelfusano (Rome, Italy) after the crown fire
occurred on July 2000.
A damage severity map was elaborated interpreting NIR false color and NDVI images derived by a Ikonos
image acquired on the 20th of August 2000, integrated by GPS based ground survey. The following damage
classes have been mapped: no damage (class 0); damages prevalently due to surface fire, most of vegetation of
lower layers is damaged but most part of top canopy is undamaged (class 1); dominant top canopies are partially
damaged, or the fire is patchy and damaged and undamaged areas are mixed together in a mosaic of elements
smaller than the minimum mapping size of 0.5 ha (class 2); most of tree vegetation is significantly damaged by
crown fire (class 3).
To support post-fire reclamation activities, a second map (precision logging map) was created pan-sharpening
the NIR false color images elaborated by the composition of the bands 4, 3 and 2 (Figure 4). The image had the
resolution of the panchromatic channel (1 m), and it could be effectively overlaid to forest road and timber
yarding network for supporting harvesting planning to remove burnt and partially unburnt material and to secure
the area (Chirici et al. 2001).
63
Gherardo Chirici
Figure 4: Pansharpened near-infrared Ikonos image acquired after the forest fire in the woodland of
Castelfusano (Rome, Italy). Burnt areas are in green and black. Unburnt areas are represented by the red
canopies, characterized by high photosynthetic activity.
3.3. Field assessment and attribute spazialization
The restoration of forest and other wooded land damaged by wildfires should be designed through the analysis of
vegetation syndynamics and dendrostructural features of the residual stands.
For such tasks, the implementation of a network of field permanent geocoded sample plots is required to follow
post-fire natural dynamics and regrowth effectiveness. The use of several small sample plots, of size around 1020 m2 in burnt forest stands, is usually better than a few large plots, at least considering the monitoring just after
the fire (Corona et al. 1998). Attributes to be measured are, for instance: deadwood; tree natural regeneration
(sexual and vegetative); vegetation diversity.
Field inventoried data may be spatialized in order to produce map of the main interesting attributes, using
parametric or non-parametric statistical tools, and, preferably, integrating independent information from
remotely sensed images (e.g., through multiregression or k-Nearest Neighbour procedures).
An example of the application of a parametric correlation method to describe the post-fire broadleaves tree
regeneration in the Castelfusano forest is presented in Figure 5. Post-fire natural regeneration mapping was
carried out integrating GPS geocoded field data collected in 178 rectangular transects of 20 m2 each with spectral
data from a QuickBird image. Information acquired in the field were related to spectral data of the QuickBird
image and to soil types in order to establish a statistical relationships for spatializing a tree regeneration index. A
multiple linear stepwise regression procedure was succesfully applied: band 4 and SAVI (Soil Adjusted
Vegetation Index) from QuickBird imagery resulted as the most significant predictors of the broadleaf sexual
regeneration in the considered area.
64
Figure 5: Broadleaves natural regeneration index (gamic regeneration) spatialised on the basis of a
multiregression analysis by QuickBird post-fire image and field data in the Castelfusano woodland
(Rome, Italy).
4. Selected ongoing applied developments
Research in the field of remote sensing technology applied to support forest fire fighting (risk prediction, early
detection) and monitoring has been rather active in the last decades. Most of the efforts are focused on large
fires, as those carried out at EC-JRC leading to the mentioned EFFIS RDA service. However, it is deemed
interesting to give here some selected hints on other ongoing applied research programs: more detailed
information can be found on many Internet websites.
European Space Agency (ESA) developed a multi-year global fire atlas (ATSR World Fire Atlas) starting in the
1995 and based on data acquired by the Along Track Scanning Radiometer (ATSR) sensor on ERS-2 satellite,
launched in 1995, and the Advanced Along Track Scanning Radiometer (AATSR) sensor on Envisat satellite,
launched in 2002. Burning fires are classified measuring thermal infrared radiation at night by ATSR/AATSR
sensors. These data are available to registered users online in near-real time (ESA asserts approximately six
hours after the acquisition) (Figure 6).
65
Gherardo Chirici
Figure 6: ATSR World Fire Atlas. Number of fires occurring monthly for France, Italy, Spain, Portugal
and Greece from July 1996 to August 2007. Source: European Space Agency, 2007. [online: september
2007] URL: http://www.esa.int/images/stats_H.jpg.
As part of the GMES Service Elements Initiative of ESA, the Risk-Eos team defines, develops and delivers a set
of experimental services for flood and fire risks management, utilizing the capabilities of satellite observation in
combination with other data sources and models: the target is to map burnt areas and monitoring of vegetation
evolution with decametric resolution (minimum mapping unit = 1 ha). This service is currently delivered over
various regions of Spain, France and Italy, in particular burn scar mapping is delivered in Basilicata, Sicilia and
Sardegna region covering an area of 60000 km2. Complement to burn scars mapping is the rapid mapping
products for large fires to support the International Charter for major events and local collectivities.
Potentially relevant activities are those by the COSMO-SkyMed project developed under the responsibility of
the Italian Space Agency. The service is based on a minisatellite constellation (4 medium-size satellites),
equipped with a microwave high resolution synthetic aperture radar (SAR) operating in X-band, designed to pick
up data of environmental interest with frequent revisiting time. The system will be devoted to the civil
protection, environmental and climate monitoring, prevention of the catastrophes, control of the coasts, soil
conservation, etc., mainly for the area of the Mediterranean (prevention, alarm, crisis management, post-crisis
tools). The first satellite was successfully launched on June 2007, while the launch of the other satellites will be
completed within the early 2009.
In the United States, fire fighting strategy is based on the idea of concentrating all the possible efforts to
implement a unique coordinated service. The Geological Service coordinates the development of the Wildland
Fire Support, an information service devoted to collect all data useful for prevention, alarm and monitoring of
wild fires. The service is available on-line (www.geomac.gov) both for the citizens interested in having
information of forest fires and for fire fighters who may also upload data from field monitoring. Near-real time
monitoring programme, based on Modis and Noaa-Avhrr data, acquires daily images to identify hot spots and to
monitor large events. Mapping activities are prevalently oriented to large fire monitoring because fire events in
the USA cover frequently very large wild areas of difficult accessibility.
5. Conclusion
The last International Wildland Fire Conference held in Sevilla (Spain, May 2007) has once more stressed the
importance of proper strategies for post-fire management based on sound monitoring, not only to pinpoint the
66
actual consequences of wildfires on land cover, land use and the environment (climate included) but also to
support effective interventions fostering ecological and economical restoration of the struck territorial systems.
This note has highlighted the potential of consistent remote sensing tools for surveying and monitoring burnt
forest areas. However, applied research is still required, distinctively to develop and refine procedures for
monitoring small fires (smaller than 10 ha) over large territories, like those typical of most European countries
(Mediterranean ones included). Moreover, the assessment of forest burnt area is not enough: the development of
reliable tools for the quantitative assessment of forest damages, biomass loss and carbon emissions due to
wildfires is to be significantly advanced.
Futhermore, from a practical side, future research efforts needs to be oriented towards an effective integration of
available data sources. For instance, for the implementation of a national system for mapping forest burnt stands
at 1:10000 scale in a country like Italy or Spain, the most effective current scenario might be the modular
integration of GPS field surveys with targeted acquisitions (with respect to areas difficult to be field-surveyed
because of low accessibility or GPS signal lacking or in the case of very wide and/or very patchy fires) of
multispectral imagery by aerial light vectors or satellites carrying VHR sensors, even eventually coupled with
targeted acquisitions (one at the beginning and one at the end of the fire season) by medium resolution satellite
imagery on larger scales.
Acknowledgements
The work, carried out by the Autors in equal parts, has been partially funded by the Italian Ministry for Research
and University under the project CarboItaly (FISR funds). The Authors thanks Agriconsulting S.p.A. for
allowing the use of data presented at the § 2.3.1.
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68
PERMANENCE OF THE CARBON STOCKS IN THE NORTH AMERICAN
BOREAL FOREST UNDER NATURAL AND ANTHROPOGENIC DISTURBANCE
REGIMES
Jean-Francois BOUCHER ,
Simon GABOURY, Réjean GAGNON, Daniel LORD, and Claude VILLENEUVE
Université du Québec a Chicoutimi - Département des Sciences fondamentales
[email protected]
Jean-Francois Boucher, Simon Gaboury, Réjean Gagnon, Daniel Lord, And Claude Villeneuve
PERMANENCE OF THE CARBON STOCKS IN THE NORTH AMERICAN BOREAL FOREST
UNDER NATURAL AND ANTHROPOGENIC DISTURBANCE REGIMES
General overview of the Canadian boreal forest
The boreal forest is one of the most important continental ecosystems, covering 14.5% of the world continents
(13.7 M km2) and containing a large fraction of the planet’s terrestrial C. The boreal forest contains 26% of the
terrestrial C stocks worldwide, and 31% of the C contained in all forest soils in the world (Melillo 1993, Dixon
et al. 1994, Gower et al. 1997). The Canadian boreal forest, with 295 Mha, represents ca. 22% of the world’s
boreal forest.
Major tree species of the Canadian boreal forest are five (5) coniferous species – Black spruce (Picea mariana
[Mill.] B.S.P.), White spruce (Picea glauca [Moench] Voss), Jack pine (Pinus banksiana Lamb.), Balsam fir
(Abies balsamea [L.] Mill.), Tamarack (Larix laricina [Du Roi] K. Koch) – and two (2) broadleaves species –
Trembling aspen (Populus tremuloides Michx.), and White birch (Betula papyrifera Marsh.) (Rowe 1972).
Overall, the boreal forest of Canada is for the most part coniferous (Figure 1). Black spruce is clearly the
dominant tree species in Canada’s boreal forest and often forms in the East mono-specific black sprucefeathermoss closed-crown stands, while mixed stands of Black spruce with Jack pine, Balsam fir, or White birch
are more common through the West of Canada (Rowe 1972). Typically, open Black spruce-lichen woodlands are
increasingly widespread from south to north of the boreal forest, as this stand type is general in the northernmost
subzone called the Taïga (Rowe 1972, Payette 1992).
Figure 1. Major forest types in Canada. The coniferous (softwood) type is in dark green.
Source: Natural Resources Canada (2001).
The managed forest of Canada covers 240 million ha (over the 310 ha of total forest land), most of which (93%)
is publicly owned. As most of the population lives outside of the forest land, the forest structure, composition,
and age is essentially determined by the physical environment (climate variables, soils, orography, etc.) and
major disturbances, both natural and anthropogenic (e.g. Bond-Lamberty et al. 2007). Major disturbance agents
in the Canadian boreal forest are of three kinds, namely: harvesting, wildfire and insect outbreaks. Two insects
are responsible of most of forest area affected historically, the Mountain pine beetle (Dendroctonus ponderosae
Hopkins, Coleoptera: Curculionidae, Scolytinae) mainly in Western Canada, and the Spruce budworm
(Choristoneura fumiferana Clemens, Lepidoptera : Tortricidae) mainly in Eastern Canada (Kurz et al. 2008a).
The area of Canadian forest harvested each year is on average (years 1990-2005) 2.3 times less of that burned by
fire, and 17 times of that defoliated and/or killed by insects (Fig. 2).
70
18
45
16
40
14
35
12
30
Area (Mha)
10
8
25
20
15
6
10
4
5
2
2005
2004
2003
2002
2001
2000
1999
1998
1997
1996
1995
Harvest
1994
Fire
1993
Insects
1992
-
1991
1990
Area (Mha)
Permanence of the carbon stocks in the north american boreal forest under natural and anthropogenic disturbance
regimes
Figure 2. Area of forest affected by insects (defoliated by insects and beetle-killed trees), fire, and
harvesting in Canada between 1990-2005 (over the 402 M ha of total wooded land). The averages are
presented in the graph on the left, and the yearly variations are presented in the graph on the right
(triangles = harvest, squares = fire, diamonds = insects). Source: Natural Resources Canada (2008a).
Permanence of C stocks in lands under forest management
Impacts of major disturbances on C stocks
If the boreal forest has been a net C sink in the last century (eg. Balshi et al. 2007) – in part because of increased
net primary production caused by climate change (Nemani et al. 2003) – the sink capacity is presently under
serious threat, and further more in the future with the anticipated climate change (Kurz et al. 2007 and 2008b).
Several lines of evidence point to a decrease in the sink capacity of the world forest in the future because of
climate change impacts, and positive feedbacks to climate change from the forest (Balshi et al. 2007, Kurz et al.
2007, 2008a and 2008b, Heimann and Reichstein 2008). The increase in intensity and frequency of natural
disturbances will probably not only prevent the CO2 fertilization and elevated temperature to enhance the net
ecosystem production of the forest land (Flannigan et al. 2005, Balshi et al. 2007, Kurz et al. 2007), but will also
feedback to the climate by increasing C emissions (Kurz et al. 2008a and 2008b).
In Canada, the recent Mountain pine beetle outbreak, which is of unprecedented extent and severity, will result
(at the end of the outbreak period of 2000-2020) in the cumulative emissions of 270 Mt C – caused by both the
reduction of net primary production and the increase in heterotrophic respiration – which is equivalent to approx.
5 years of emissions from Canada’s transportation sector (Kurz et al. 2008a). This single insect outbreak has
initiated a transition of the Canadian forest from a small net C sink to a rather large net C source, both during and
immediately after the outbreak (Kurz et al. 2008a). This has also recently contributed in Canada’s decision not to
elect forest management under the Article 3.4 of the Kyoto Protocol (Figure 3), since this would have further
separate Canada from its reduction target in emissions for the first commitment period (Natural Resources
Canada 2007, Kurz et al. 2008b).
Figure 3. The managed forest sink and source in Canada from 1990 to 2005. Note that the appearing
discrepancy in area disturbed between this graph and the graph presented in Fig. 2 is because of the
larger area of reference in the latter graph. Source: Natural Resources Canada 2007.
71
The conservationism issue
Environmentalists have recently suggested that forest management in Canada is a threat for the climate, as it
might activate the “carbon bomb” held within the natural boreal forest (Ferguson et al. 2008). They then claim
for the conservation of the natural boreal forest for the sake of mitigating climate change. Most of this
argumentation is based on the idea that forest harvesting contributes to GHG emissions by releasing some of the
immense C stocks locked in the intact boreal forest, and that unmanaged forest is better suited to face climate
change than managed forest. These assumptions are misleading on at least three points. First, the temporal scale,
ie. if it is true that on the short term forest harvesting corresponds to a net emission compared to the baseline
scenario, on the long term the harvest disturbance will approach zero emission since the forest growth will regain
the initial lost, like any other self replacing forest after a natural disturbance. Secondly, these assumptions ignore
the contribution of logging to the permanence of C stock in the wood products, or the production of bioenergy,
especially when the wood products are used instead of energy-intensive materials, like steal or concrete, or when
the bioenergy is used instead of fossil fuel (Kurz et al. 1998, Baral and Guha 2004, Perez-Garcia et al. 2005,
Fleming et al. 2006, Kirschbaum 2006, Olsson and Kjällstrand 2006). Thirdly, forest management can contribute
to the climate change mitigation portfolios, particularly by increasing harvest rotation lengths, reducing
regeneration delays, or increasing stocking densities (Nabuurs et al. 2007, Kurz et al. 2008b). Even if this
contribution to GHG removals might not compensate the net emissions associated to the impacts of natural
disturbances (Kurz et al. 1998, Balshi et al. 2007, Kurz et al. 2007, 2008a and 2008b), it could help impeding the
positive feedback to climate change caused by natural disturbances on either managed or unmanaged forests
(Nabuurs et al. 2007, Heimann and Reichstein 2008, Kurz et al. 2008b).
Permanence of C stocks in afforested lands
Human induced vs natural deforestation
Deforestation accounts for up to 20% of the global anthropogenic emissions of GHG (Nabuurs et al. 2007).
About half of the 1.3 to 4.2 Gt CO2eq estimated mitigation potential per year by 2030 (at carbon prices < 100
US$/tCO2eq) can be achieved by reducing emissions by deforestation, most of which in the tropics (Nabuurs et
al. 2007). Part of the remaining human induced deforestation (ie. land-use change) is permanent and is caused by
expansion of settlements and infrastructure, while other parts are less permanent and are suitable for
afforestation/reforestation, for example on abandoned agricultural lands or fallow lands (Nabuurs et al. 2007). In
Canada, deforestation affected les than 0.02% of the Canada’s forests in 2005, and accounts for less than 3% of
Canada’s total GHG emissions and 0.4% of the global deforestation (Natural Resources Canada 2008b). The
economic potential of CO2 sequestration by reforestation in Canada would be 50 to 70 Mt CO2/year, and approx.
7.5 Mha of agricultural land could be economically attractive for poplar plantations (McKenney et al. 2004,
Nabuurs et al. 2007).
These figures, however, do not account for another type of deforestation, which appears to be unique to the
boreal forest: the natural deforestation of black spruce (Picea mariana (Mill.) B.S.P.) forests. In Canada’s
Eastern boreal zone, the black spruce-feathermoss (BSFM) domain (between the 49th and the 52th parallels)
covers 28% of the province of Québec. Black spruce is the dominant tree species, representing more than 75% of
the forest cover in the BSFM domain (Bergeron 1996, Gagnon et Morin 2001). While black spruce is generally
well adapted to regenerate after wildfire (Heinselman 1981, Viereck et Johnston 1990), regeneration accidents
can sometimes occur, resulting in the irreversible conversion of closed-crown BSFM to open black spruce-lichen
woodlands (hereafter shortened to open woodlands or OW) (Payette 1992, Gagnon et Morin 2001, Jasinsky and
Payette 2005). There is presently no evidence of natural redensification of OW, i.e. a shifting to a closed-crown
BSFM stand (Payette 1992, Jasinsky and Payette 2005). Moreover, a recent study showed a gradual increase in
OW creation during the last 50 years (Girard et al. 2007). The last Québec’s forest inventory reveals that
approximately 7% (1.6 M ha) of the BSFM domain is made of OW (MRNF, 3rd decennial forest inventory), of
which nearly 10 % are under 5 km of distance from the already existing road network in 2002 (Plante 2003).
The case-study of the afforestation of boreal open woodlands
Afforestation of OW have been tested recently with an experimental plantation network within Québec’s central
boreal zone, where site prepared OW were compared to adjacent and managed BSFM stands (Girard 2004,
Hébert et al. 2006). The first results from these tests of OW afforestation show significant growth and survival of
seedlings, within the establishment window of three years after plantation (Hébert et al. 2006). The large amount
of available OW within the closed-crown boreal forest in the province of Québec – and probably further more
across the boreal forest of Canada (Rowe 1972) – represents a significant theoretical potential for C
sequestration that has not been evaluated yet.
72
regimes
We have estimated the theoretical C balance of the afforestation of OW within the closed-crown BSFM domain
in Québec’s boreal forest, and calculated – using the life cycle analysis (LCA) methodology – all the GHG
emissions related to black spruce OW afforestation in the closed-crown BSFM domain of Québec (Gaboury et
al. 2008). The CO2FIX v. 3.1 model was used to calculate the biological C balance between the baseline (natural
OW of site index 9 at age 50) and the afforestation (black spruce plantation of site index 6 at age 25) scenarios,
using the best estimates available for all five recommended C compartments (above ground biomass,
belowground biomass, litter, dead woods, and soil). The simulation revealed a biological C balance of 77.0 t C
ha-1 after 70 years following afforestation, for a net sequestration rate of 1.1 t C ha-1 year-1 in average. Biological
C balance only becomes positive after 27 years (Figure 4).
120
Biological net C balance (t C ha-1)
100
Total stem
Leafs
Branches
Roots
Soil
Total
80
60
40
20
0
0
-20
10
20
30
40
50
60
70
80
Time (years)
Figure 4. Biological carbon balance of the afforestation project during 70 years. Results are from the
CO2FIX model (Masera et al. 2003; Schelhaas et al. 2004).
When integrating the uncertainties related to both the plantation growth yield and the wildfire disturbance, the
average sequestration rate varies between 0.2 and 1.9 t C ha-1 year-1 (Figure 5). GHG emissions are of 1.3 t CO2
eq ha-1 for all afforestation-related operations, which is less than 0.5% of the biological C balance after 70 years.
Thus, GHG emissions do not significantly affect the net C balance of the afforestation project simulated. In the
paper where these results are presented (Gaboury et al. 2008), we have made recommendations, mostly centered
on the factors influencing the growth rate of carbon stocks and the impact of natural disturbances, to minimize
the range of uncertainties associated to the sequestration potential and maximize the mitigation benefits of an
OW afforestation project.
Figure 5. Impacts of uncertainties related to fire and plantation wood yield on global C balance of an OW
afforestation project. The fire severity impact is from Bergeron et al. (2004).
73
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76
WINDTHROW RISK MANAGEMENT. RESULTS FROM ROMANIAN FORESTS
Ionel POPA
Forest Research and Management Institute - Campulung Moldovenesc
[email protected]
Ionel Popa
WINDTHROW RISK MANAGEMENT. RESULTS FROM ROMANIAN FORESTS
1 Introduction
The natural mountainous forest ecosystems are characterized by a high structural diversity and they impress by
their stability. In the last decades, frequent ecological disturbances have been reported in the mountainous
ecosystems with high consequences on their productivity and their capacity to satisfy the multiple functions
assigned. Promoting an environmentally friendly silviculture, based on ecological principals, preserving and
developing on sustainable tenets these complex ecosystems represent the most important problems faced by the
modern forest management.
The ecology of disturbance factors represent a new domain of basic and applicative research with major
application on the knowledge of the interaction interface between forest ecosystems and environmental
disturbance factors (wind, snow, fire, insect etc.) (Pickett and White, 1985). Disturbance factors are a natural and
integrate component of the forest ecosystems (Dale et al., 2000; Peterson, 2000). The succession processes in
forests, and mainly in mountain forests, which occur after the action of a disturbance, the ecological role of these
been well defined (Dale et al., 2000; Ulanova, 2000; Kramer et al., 2001; Popa, 2007).
Wind damages, by their negative economical and ecological effects constitute an actual problem of forestry
research which importance was recently accentuated by the necessity of adaptating management plans to the
sustainable forest development requirements (Popa, 2001).
The analysis of the structure and the function of the virgin forest ecosystems, those kept out of human
disturbance, bring to the conclusion that windthrow events are a normal process, natural and perfectly integrated
in forests biogeochemical cycles. Windthrow events can be considered as a manifestation of ecosystem functions
that constitute a form of self thinning process together with other specific processes. When analyzed at large
scale, windthrow events do not constitute a disturbing factor for ecosystems having reached climax, that is, for
which the structure stability is maximum, their rate is within the range of natural elimination rates.
In comparison with natural forests, the forest ecosystems in which human activities Windthrow damages
overcome self-thinning’s intensity in managed forests in which human activities modified the structure and the
relation between the components of the ecosystem, therefore becoming a factor of disturbance having negative
effects on both economic and ecologic functions through the modification of the stand structure and the
economic disordering it induces. The annual economic losts of this calamity are in the range of hundreds
thousands of Euros at European level. For example, in December 1999 only, the storm Lothar affected over 200
millions m3 over Europe. These negative effects are sensed at Romanian level too, particularly with the
catastrophic damages of November 1995 with over 140,000 ha affected and a total volume of about 7.9 million
m3, while the damages that occurred in March 2002 in Suceava County where the woodblown volume was
estimated to 6-7 million m3.
The knowledge on realistic basis of the complex processes represented by wind damage and the implementation
of the windthrow risk management systems, in the condition that the mountain forestry is a silviculture in risk
conditions, represent a key to assure the sustainable forest management.
2 Impact of wind damage on forests
The disturbance is defined like a relative event, discrete in time and space, which induce an alteration of the
structure of the population, community or ecosystem by the modification of the resources accessibility or
physical environment (Pickett and White, 1985). Windthrow represent one of the main disturbance factors of the
mountain forest ecosystems (Ulanova, 2000). By the term of windthrow or wind damage we understand a
mechanical damage to a tree or stand as follow of the wind action. The wind disturbance varies as well spatially
as temporally in forest ecosystems, from catastrophic damage at the level of landscape to local disturbance at the
level of tree (Picket and White, 1985; Kuuluvainen, 1994). In rapport with the nature of the damage induced by
wind we can distinguish structural damages (uprooting, stem breakage etc.) known as windthrow and functional
damages such as the breakage of fine roots, the modification of water balance by increase of transpiration.
For a better understanding and analysis of the wind damage effects, we adopt a classification of the disturbance
as: catastrophic windthrow and endemic windthrow, having as criteria the intensity and the spatial extend of the
event at a specified analysis scale (Miller, 1985).
The catastrophic wind damage are the massive disturbance induced by specific weather conditions (high speed
wind), which affects large forest surface. The delimitation criteria is the total volume of damage wood which can
be in the order of hundred of thousand cubic meter or million cubic meters in rapport with the geographical scale
of analysis. For example, at European level, the minimum intensity at one event can be 1 million cubic meters,
and at the level of Romania we can adopt the level of 100.000 m3. To include in this category a wind damage
event some specific conditions must be reached:
- the surface affected by wind in the order of hundred of hectares on a relative small geographical space;
- the volume of wood is in the order of hundred of thousand m3, the majority of them been from stands
completely blown down;
78
Windthrow risk management. Results from romanian forests
it is determined by particular extreme weather conditions;
- the local topography has a major role in the risk mapping of these catastrophic events;
These wind damage are relatively rare, every 10-15 years, and affect the entire stand indifferently of the structure
or their stability. Modeling catastrophic windthrow events has two principal components:
- the estimation of the occurrence probability by statistical methods;
- establishment of risk maps using simulation on digital terrain models of the speed and directions of
wind;
Endemic windthrow are the wind damages recorded every year in mountain forests determined by wind with
medium speed. They have the highest cumulative economic effects. Contrary to catastrophic windthrow, these
types of disturbance affect large areas with different intensity. They are induced by topographical and weather
factors as well as stand characteristics (biometric parameters, health status, stand structure etc.). To quantify the
endemic wind damages, divers indicators are proposed in literature. To achieve a better understanding of the
influence factors we propose for the quantification of endemic windthrow the follows parameters:
- the occurrence probability of endemic windthrow (p) calculated as ratio between the number of events
(damaged stand or year with wind damage) and the total number of studied cases;
- the percent of windthrow (Pdv) to quantify the intensity of disturbance, computed as percent of the
damages wood volume from the total stand volume before the action of the disturbance factor.
In the processes of economical management of risks, it is recommended to use indications with economical
basis, respectively to convert the synthetic indicators, presented above, in economic units.
2.1 Chronology of catastrophic windthrow
The forests of Romania, like other forests from Europe, were affected in the last years by catastrophic wind
damage (Ichim, 1990; Popa, 2001). The first references about wind damage for Romania are from 1885, 1905,
1915 and 1916 when are reported important windthrow in Eastern Carpathians (Marcu et al., 1969). Between
1829 and 2005 are recorded a total number or 34 catastrophic windthrow, the most significant been from
1947/1948, 1964, 1969, 1973, 1995 and 2002. From the chronology of catastrophic wind damages for Romania
we observe an increase of the frequencies and intensity in the last decades (fig. 1). Norway spruce has the
highest percentage amongst species with over 85% of the total damaged wood volume.
8000
Thousands m3
6000
4000
2000
0
1880
1900
1920
1940
1960
1980
2000
Year
Fig. 1 Chronology of catastrophic windthrow from Romania
The actual dynamics of cumulate wind damage volume has an exponential trend, while three periods can be
distinguished:
79
Ionel Popa
- the first one that lasted up to 1947-1948 characterized by a relatively low frequency and intensity of the
phenomenon. These could result from a lack of exact records, the forest service being doing its first steps at that
time, or could result from the predominance of the resistant natural coniferous forests while the stands artificial
and uniform were young and covering a limited area.
- the second period between 1947-1948 and 1974-1975 when the strongest windthrows occurred. The lost were
counted in millions of m3, with a windblown volume over 30 million m3 within a quarter of a century only. The
spruce stands in Bucovina were particularly concerned, perhaps because of an intensified wind activity during
this period. A more likely explanation is that the monospecific stands generated at the beginning of the century,
under the impulse of the German forestry school, reached their maximum sensitivity at that time.
- the last period from 1975 during which can be observed a reduction of both frequency and intensity of the
phenomenon.
It seems that we entered in the last years in a new period with intense windthrow, the most recent being that of
March 2002 (over 7 millions m3) and March 2007 (more that 1 million m3) both of them in Suceava county. The
frequency of catastrophic windthrow by decades is presented in figure 2.
7
6
No. of windthrow
5
4
3
2
1
0
1880 –
1890
1891 –
1900
1901 –
1910
1911 –
1920
1921 –
1930
1931 –
1940
1941 –
1950
1951 –
1960
1961 –
1970
1971 –
1980
1981 –
1990
1991 –
2005
Period
Fig. 2 Frequency of catastrophic windthrow in Romania by decades
The economic and ecological effects of these catastrophes are significant, inducing important disturbance of
local and regional wood market, in the implementation of the management plans. The modeling of this type of
disturbance it is difficult but some statistical techniques of extreme value can be applied.
2.2 Effect of endemic wind damage on management plans
The endemic wind damage, unlike catastrophic windthrow, do not impress by intensity of affected surface, but
represent the most important disturbance factor of the mountain silviculture from Carpathians (Popa, 2007). The
catastrophic windthrow determine a high impact on the media and complex action plans are implemented, the
endemic wind damage remain unobserved, become a normal event for most of the forest districts from mountain
region. The economic analysis of this disturbance on the management plans would permit a reconsideration of
the place and importance of endemic windthrow.
The vertical analysis, from a low to a high level of organization enables the identification of potential risk zones
for each organization level. A decision factor can then be obtained concerning the frequency and the intensity of
the phenomenon therefore enabling an efficient allocation of the human and economic resources. The
chronologic series concerning the windthrow events together with the researches carried on in our country in this
domain offer a mean to situate the forest ecosystems of Bucovina, in the context of its integration in the
Romanian forest found, in a zone with particular vulnerability to windthrow. As being a high-level risk zone, the
macro zonal situation urges the adoption of measures at intermediate and micro-zonal level to differentiate the
80
existing ecosystem complex into risk classes with regards to windthrow because all stands in Bucovina do not
have a high level of wind sensitivity.
Applying this spatio-temporal method to the level of a forest management district, considering the district as the
references, it is possible to identify the zones of potential risk, but at a lower organization level. For example,
this methodology has been applied to the forest found of Suceava’s county, managed by the forest directorate of
Suceava, located in a zone of high risk according to the macro zonal level determined. Primary data consist of
the windthrow volume entering the economic circuit for the period 1981-2000, without catastrophic windthrow.
The ratio of the wind damages volume to the total harvested volume at the level of the forest directorate of
Suceava is very high, ranging between 10 to 60% and exceeding 500,000 m3 on average (fig. 3). We observe the
existence of a period of minimum occurred between 1984-1989 and 1991-1992, while at present it is located
above 50% of total volume. At decadal time-step, the ratio of accidental produces increased from 26% for the
period 1981-1990 to 46% for the decade 1991-2000, and respectively from 30 to 54% dealing with conifers. The
percentage of windblown volume is much greater in this case reaching 60-65% in the last years.
2500
Thousands m
3
2000
1500
1000
1994
1995
1996
1997
1998
1999
2000
1995
1996
1997
1998
1999
2000
1993
1994
Windthrow cuttings
1992
1991
1990
1989
1988
1987
1986
1985
1984
1983
80%
1982
0
100%
1981
500
Other cuttings
60%
40%
20%
Windtrow cuttings
1993
1992
1991
1990
1989
1988
1987
1986
1985
1984
1983
1982
1981
0%
Other cuttings
Fig. 3 Dynamics of wood volume from windthrow in Suceava directorate
This increase of the ratio of wood from windthrow harvested at the scale of the Suceava’s forest directorate
results from the unbalance between a significant reduction after 1990 of wood harvested from final cuttings and
the slight increase in the volume of accidental produces. The volume of wind damages is much higher than the
one of normal harvest wood, being twofold during the period 1993-1996, with the exception of years 1985-1988.
Analyzing the time-dynamic of the accidental produces volume harvested from stands over 60 years old, as
compared to final harvest produces volume; one can observe that negative impacts are higher. For the period
1993-1996, the windblown wood volume was three times greater than the volume harvested by application of the
silvicultural treatment. Practically, half of the regeneration cuttings that were programmed in the decadal plan
were not concretized if counted that those accidental produces are retained from the final harvest potential
volume.
The dynamic of the harvests in Suceava’s forest direction over a 20 years period shows that windblown wood
represents a significant proportion of harvested wood volume, but the fact that they are not accounted for in the
forest management and in the computation of the possibility indicators as well as the estimation of an accidental
volume equal to zero proves to be unrealistic. The dynamic of the annual volume harvested from windthrow per
hectare ranges between 3 and 48 m3·ha-1·an-1, depending on the surface on which accidental cuttings were
done. Dividing the total wind damages harvested volume to the surface of the Suceava’s forest direction, one can
observe that the average index of accidental produces harvested vary between 0.4 m3·ha-1·an-1 to 1.8 m3·ha1·an-1. On average, they reach 1 m3·ha-1·an-1 which represents a lot if compared to the total forested area of the
county of Suceava that covers 410,000 ha. While indexed to the total coniferous area, the accidental produces
represent from 0.5 m3·ha-1·an-1 to 2.1 m3·ha-1·an-1, with an average of 1.14 m3·ha-1·an-1. Forest districts of
the western and north-western side of Suceava’s forest directorate, that is Cârlibaba, Iacobeni, Coşna, Dorna
Candrenilor, Panaci, Brodina, Falcău and Moldoviţa show a high risk with an average rate of windthrow above
1.0 m3·ha-1·an-1, which represents around 25% of the annual current growth(fig. 4,5).
81
Ionel Popa
Fig. 4 Spatial distribution of mean percent of endemic windthrow in Suceava directorate
200
Thousands m
3
O.S. Cârlibaba
150
100
50
1991
1992
1993
1994
1995
1996
1997
1998
1991
1992
1993
1994
1995
1996
1997
1998
1991
1992
1993
1994
1995
1996
1997
1998
1990
1990
O.S. Dorna Candrenilor
Windthrow cuttings
1990
1989
Alte produse
1988
1987
1986
1985
1984
1983
1989
1988
1987
1986
1985
1984
1983
1982
Alte produse
Produse accidentale
1982
1989
1988
1987
1986
1985
1984
1983
1981
1982
1981
120
100
80
60
40
20
0
O.S. Iacobeni
Produse accidentale
1981
Thousands m
3
Thousands m
3
0
120
100
80
60
40
20
0
Other cuttings
Fig. 5 Endemic windthrow in some forest district from Suceava directorate
Besides those catastrophic wind-induced events that impress by their intensity, magnitude and their economic
effects, endemic windthrow events, having a much reduced intensity but greater frequency, constitute the main
disturbance factor for mountainous forest ecosystems with long-term consequences on economy and on the
82
ecology. Forest managers cannot ignore those disturbances but should study them in order to find means to
diminish their impact. The reduction of the impact depends largely upon the management of forest resources and
the forest economy decision system.
3 Modeling of catastrophic windthrow
A detailed analysis of the classification criteria of the wind damages as presented in forest literature was done in
order to correctly interpret and classify some terms otherwise misused. The analysis of the risk of wind damages
was done differentiating between the catastrophic wind damages and the endemic wind damages.
Natural hazards, of which windthrows and high-intensity snow damages, are rare, but using the large time step
that can be encompassed in available data, it appears that statistical and mathematical-based models can be used
to interpret the data. Using the modern method of extreme value theory on catastrophic windthrows at national
level, it has been possible to do a probabilistic estimation of the windthrow damage frequency (fig. 6).
Fig. 6 Weibull model for decadal risk of catastrophic windthrow in Romania
At the Romanian level, based on a Weibull model, we can forecast that the probability of a massive wind
damage occurrence within a ten-year period is 55%, while the likelihood to have a second windthrow is 33%.
Applying the modern techniques for analysis the seasonality of windthrow, respectively the autocorrelation
function, partial autocorrelation function and spectral analysis, we found that the dynamics of catastrophic wind
damage at national level shows seasonality with a return period of 3-4 and 56 years (fig. 7).
83
Ionel Popa
Fig. 7 Spectral density function of the catastrophic windthrow chronology in Romania
4 Endemic windthrow risk assessment
Combining the knowledge provided by statistics and modeling, by doing spatial and temporal analyses of the
stability of trees, stands and forests as a whole, generates a new level of the integration of the observations and
data gathered by classical research methods. It is a necessity to understanding the windthrow implications.
The evaluation of the risk in forestry did not receive a great attention due to the complexity of the phenomenon.
An exact knowledge of the risk is however necessary, and of the factors that drive it as well as the extent of the
lost resulting. In the absence of an objective evaluation of the risk, the decision factors of the forest management
and sector-based policy only have a post factum answer imposed before all by the level of economic lost. This
reaction is often disproportioned as compared to the risk level. The reaction of the forest decision factors is fast,
extensive and complex in the case of a catastrophic event that would impress by the volume concerned. But there
are no reactions in the case of an endemic damage that would have a much more reduced intensity compared to a
catastrophic event and therefore have a much greater generalized economic impact. An eloquent example is
constituted by the windthrow of March 2002 that occurred in Suceava’s forest directorate and induced a more
complex mobilization of the forest managers as compared to the endemic damages that occur every year.
The evaluation and the quantification of windthrow damages requires an integration of the biometric and
structural characteristics of the stands and the trees within a statistic model of risk evaluation, thus providing a
system of the risk induced by the vegetation. This component of the risk brings information concerning the
management of the vulnerability to windthrow over short time-steps, thus enabling the construction of a strategy
for the period of a production cycle. Through the probabilistic combination of those two types of risk evaluation
methods, i.e. topographic and vegetation-induced, a picture of the current vulnerability of the forest ecosystems
of the region studied can be obtained.
The theoretic stands stability to wind action is extremely complex. The interactions between the factors that get
involved in the stability process of the stands are very complex, with the consequence that their mathematic
quantification is extremely difficult with the actual mathematic tools. The great variability of the factors that are
involved in the system imposes the adoption and the construction of probabilistic models (statistical models) for
analyzing and forecasting windthrow events at short and long term. The scope of modeling the stability of stands
is to offer an adequate and efficient instrument for measuring the susceptibility of a station and a stand to the risk
of a windthrow occurrence, and also for estimating the likehood of future perturbations.
The spatial and temporal analysis of wind damage, also of the complexity of the factors involved have made
possible a synthesis of the windthrow likehood in a probabilistic relation composed by the probability of
84
appearance of wind damage (p) and the percentage of windthrow (Pdv) expressed as the percent of wind damage
volume divided by the volume existent per ha.
The evaluation and the quantification of hazard exposition find a direct output in the mapping of the topographic
risk. In the case of windthrow events, the hazard exposition is mostly driven by the regional climate, the relief
being the principal factor of modification of the direction and the velocity of air streams, the topographic
exposition, the slope and the soil conditions. By integrating these elements in a statistical model, one can obtain
a map of the natural risk resulting from the topography. This system of natural potential topographic risk
mapping offers information to decision factor in the context of the elaboration and the implementation of the
forest strategies on long term for a given forest ecosystem (Fig. 8).
Fig. 8 Topographical wind damage risk evaluate by TOPEX in the basin of Bistrita
The windthrow occurrence likehood was quantified by statistic models with two factors, respectively the age and
the stand density. In the case of production units I Demacuşa we realized the stratification of the model in
rapport with the percentage of spruce in the stand composition. In conformity to models predictions, in the case
of the forest compartment VII Izvoarele Bistriţei of the district of Borşa, we observed that the stands of the age
class V-VII and having a reduced consistency present a risk over 50% to suffer from a windthrow event (fig. 9).
The variation of the annual windthrow occurrence likehood as estimated by the statistic model for the
compartment I Demacuşa of district Tomnatic, obtained per category of spruce contribution to the stand reveal
the following conclusions (fig. 10):
- in stands composed exclusively or mainly of spruce, the risk increase is driven by the age, the consistency
having a minor role.
- for mixed stands in which spruce represents 60 to 80%, a zone of high risk is identified for stand with degraded
consistency regardless of the age, and another pole vulnerable with stands over 120 years.
- a similar situation can be observed for stands in which spruce represents less than 60% of the composition,
when the consistency is reduced.
We proposed a model of wind damage percentage based on the general form of the logit model and modified by
the introduction of a correction factor k, the factor k being determined after the quantification of the statistic
parameters. In the case of the production compartment VII Izvoarele Bistriţei the stability parameters adopted,
after the regressive analysis, are in order: the exposition, the altitude, the age, the type of structure, the stand
density, the height, the class of production and the rapport height/diameter. For the production compartment I
Demacuşa the stability factors included in the model are: the slope, the altitude, the spruce percentage, the stand
85
Ionel Popa
density, the age, the height and the rapport height / diameter.
The final scope of the modeling of tree and stands stability is to obtain efficient and realistic criteria for mapping
the windthrow risk. The necessity to realize and implement a differentiated forestry technique system based on a
realistic and efficient risk mapping system is stressed by the perspective of windthrow intensification. The
current system of risk mapping, promoted by the technical rules, is too general and do not offer an efficient tool
of analysis and decision to the users.
Taking into account the financial difficulties that appear while doing thinning, along with the very low demand
for low-size wood on local place, it appears that the identification as precise as possible of the stands in which
particular measures of consolidation are required (more intensive selective thinning). Those measures are
demanding a high financial effort. This is why a system to windthrow risk mapping for zones of homogeneous
high air streams, with general vegetation conditions, etc., represents an urgent requisite measure for the
management of mountain forests.
In the spirit of these modern ideas were elaborated and implemented the following windthrow risk mapping
systems using as final indicator the likely windthrow volume expressed per area and per time unit (fig. 9, 10).
A
B
C
Fig. 9 Endemic windthrow risk mapping for forest compartment Izvoarele Bistriţei (A: wind damage
occurrence probability; B: endemic windthrow percentage; C: integrate windthrow intensity in
m3/ha/year)
A
B
C
Fig. 10 Endemic windthrow risk mapping for forest compartment Demacusa (A: wind damage occurrence
probability; B: endemic windthrow percentage; C: integrate windthrow intensity in m3/ha/year)
The windthrow risk mapping systems with annual resolution offers detailed information about the annual risk of
windthrow event, however, the method of quantification requires at least 30 years of primary data. The important
volume of data that has to be gathered and used in the statistic model and the existence of wrong records raise
some technical problems in the application of this type of model. This system of annual windthrow risk
prognosis would preferentially be applied on areas with high vulnerability and where it historic and trustable
data of wood volume already exist. A more simple system of risk mapping over the period of management plan
is required, available at decadal time-step, with a sufficient precision and that would estimate the potential
86
volume of windblown wood. The model probabilistic based is proposed in this work to estimate the likely
windblown volume at decadal time step.
The current management plans are based on future windblown volume equal to zero, which is not conform to
neither field or silvicultural practices. Similarly to sanitary wood volume that are foreseen in management plans,
one could mention a ‘possibility’ of accidental material relating to windblown material for the decade in which
the plan was designed (under the condition that the accidental produces are assimilated). Practically, the use of
the decadal model is proposed in order to realize the estimation of the windthrow risk, the quantification being
done for each zone that can be considered as homogenous dealing with windthrow exposition and topography,
using data over at least 30 years with age as single factor.
n
Vdob = ∑ pi ⋅ pdvi ⋅ Vi ⋅ S i
i =1
Several steps are needed for the utilization of such system (table 1):
- the quantification of the decadal probability of windthrow occurrence as deduced from the division between the
management units affected and total number of stands, the analysis being done by age classes and decades;
- the quantification of the risk intensity by computing the percent of windthrow damages per age class as the
ratio between total volume and the ratio between the total volume of accidental material and the total standing
volume of the corresponding age class;
- the quantification of the average decadal values per age class of the probabilistic model components;
- the forecast of the likely volume of windblown material for the decade.
Table 1 Example of quantification of probable wind damage volume - U.P. II Şesuri
Specification
Age classes
0-20 20-40 40-60 60-80 80-100 >100
Decadal probability of windthrow occurrence
1968-1979
0,53
0,53
0,67
0,65
0,50
1979-1989
0,67
0,63
0,63
0,62
0,57
1989-1999
0,34
0,60
0,69
0,68
0,41
pi
0,51
0,59
0,66
0,65
0,49
Decadal intensity of windthrow
1968-1979
0,20
0,18
0,24
0,21
0,14
1979-1989
0,15
0,18
0,22
0,21
0,27
1989-1999
0,20
0,11
0,12
0,16
0,23
pdvi
0,18
0,16
0,19
0,19
0,21
Volume and surface by age classes
Vi (m3·ha-1)
167
335
325
409
327
Si (ha)
896,9 872,3 375,0
859,5 719,0
Decadal forescast windthrow volume
Vdob/ha (m3·ha-1·dec-1)
15,7
30,8
41,7
51,4
34,4
Vdob (m3·dec-1)
14096 26858 15629 44176 24744
Today, forest management rules in Romania do not take wind damage volume into account when computing the
total allowable harvest wood volume in the management plans. Windthrows are not quantified and forecasted for
the next management period. It is considered only at descriptive level.
The current experience in forest management denies the hypothesis of a risk equal to zero, which is however the
case in current management plans. Mountain forestry represents a management under risk threat, but the
management of the risk constitutes a matter of value (economic, social or ecologic). Wind is one of the main
forest managers in the study area and the actual management plans are under pressure of wind damage.
The integration of the risk management at the level of the politics and the regulation of the forest sector and in
the operation management plans constitutes a necessity, a provocation for the mountain forestry of Carpathian
domain.
References
Pickett, S.T.A., White, P.S. (eds.), 1985. The ecology of natural disturbance and patch dynamics. Academic
Press. Orlando. 1-13.
Peterson, C.J., 2000. Catastrophic wind damage to North American forests and the potential impact of climate
change. The Science of the Total Environment. 262:287-311.
87
Ionel Popa
Dale, V.H., Joyce, L.A., McNulty, S., Neilson, R.P., 2000. The interplay between climate change, forest, and
disturbance. The Science of the Environment. 262: 201-204.
Kramer, M.G., Hansen, A. J., Taper, M.L., Kissinger, E.J., 2001. Abiotic control son long-term windthrow
disturbance and temperate rain forest dynamics in Southeast Alaska. Ecology. 82:2749-2768.
Ulanova, N.G., 2000. The effects of windthrow on forests at different spatial scales: a review. Forest Ecology
and Management. 135:155-167.
Popa, I., 2007. Managementul riscului la doborâturi produse de vânt. Editura Silvică. Bucuresti. 235 pp.
Popa, I., 2001. Modele de stabilitate la acţiunea vântului pentru arbori şi arborete. Teză de doctorat.
Universitatea Ştefan cel Mare, Suceava, 309 p.
Kuuluvainen, T., 1994. Gap disturbance, ground microtopography, and the regeneration dynamics of boreal
coniferous forests in Finland: a review. Annales Zoologici Fennici. 31:35-51.
Miller, K.F., 1985. Windthrow hazard classification. Forestry Commission. Leaflet 85. p. 3-14.
Ichim, R., 1990. Gospodărirea raţională pe baze ecologice a pădurilor de molid. Editura Ceres. Bucureşti. 186
pp.
Marcu, G., Stoica, C., Dissescu, R., 1969. Doborâturile produse de vânt în anii 1964-1966 în pădurile din
România. Editura Agrosilvică. Bucureşti. 224 pp.
88
REDUCING EMISSIONS FROM DEFORESTATION IN DEVELOPING
COUNTRIES: THE NEW CHALLENGE FOR CLIMATE MITIGATION
Giacomo GRASSI
Institute for Environment and Sustainability, Joint Research Centre of the European
Commission, I-21020 Ispra (VA), Italy
[email protected]
Giacomo Grassi
REDUCING EMISSIONS FROM DEFORESTATION IN DEVELOPING COUNTRIES: THE NEW
CHALLENGE FOR CLIMATE MITIGATION
Abstract
Tropical deforestation accounts for about 20% of human-induced greenhouse gas emissions. Given the
magnitude of this process, and the need of an active involvement of developing countries in future efforts to
combat climate change, the possibility to reduce emissions from deforestation is emerging as a decisive element
of the post-Kyoto negotiations. This paper presents some relevant issues with particular attention to the need to
develop robust estimates of emissions from deforestation and set up a credible reporting mechanism.
Why forests matter in the climate change debate
Terrestrial ecosystems have always played a prominent role in writing the “climate story”. Far from being a
passive recipient of anthropogenic emissions - the carbon reservoir in the world’s forests is higher than the one
in the atmosphere - forests interact strongly with the climate, providing both positive and negative feedbacks.
Furthermore, exploitation of forests affects the climate in a number of ways (emission of greenhouse gases,
transpiration, albedo,...). This complexity links tightly the biosphere, and forests in particular, to all the main
aspects of climate change: causes, solutions and impacts. Indeed, on the one hand, the terrestrial biosphere
absorbs ~30% of the anthropogenic CO2 emissions (mainly thanks to forests), therefore contributing to the
mitigation of the climate change (IPCC 2007). On the other hand, according to FAO (2006), annual gross
deforestation is 13 million ha. Due to faster rates of forest planting in recent years, the net rate of deforestation
slightly decreased from 8.9 million ha/yr in 1990-2000 to 7.3 million ha/yr in 2000-2005. From a carbon
perspective, however, afforestation should be considered separately from deforestation due to the asymmetry of
carbon sequestration of these two processes (“slow in, fast out“). Based on these FAO numbers (i.e, data
provided by countries) or on remote sensing analyses of tropical deforestation, estimates of emissions from
deforestation range from 2.2+0.6 GtC/yr (Houghton 2003) to 1.1+0.3 GtC/yr (Achard et al. 2004), with IPCC
(2007) assuming an intermediate value of 1.6+0.9 GtC/yr. The high uncertainty of this figure is due to many
reasons, including: (1) the lack of reliable data on the area deforested and the carbon stock of forests, due to lack
of proper forest inventories in many tropical countries; (2) the different definition of forest, which in turn affect
the assessment of deforestation; (3) the difficulty to assess greenhouse gas (GHG) emissions from forest
degradation and removals from re-growth. Irrespective of this high uncertainty, deforestation remains a huge
source of emissions, and if the IPCC estimate is taken it represents globally the second single GHG source,
behind energy production, being responsible for about 20 % of human GHG emissions.
Although the approximate magnitude of this data is well know, any past effort to link these emissions to the
climate change debate in the context of United Nation Framework Convention on Climate Change (UNFCCC)
has failed. However, in recent years, an increasing consensus emerged on considering emissions from tropical
deforestation. On the one hand, several analyses (e.g. IPCC, Stern review, Word Bank) suggested reducing
deforestation as a very cost-effective way to mitigate climate change in the short term, while potentially
producing several beneficial side-effects (sustainable development, biodiversity). On the other hand, the issue of
tropical deforestation started to assume a strategic role to involve developing countries in future efforts to
mitigate climate change.
From Kyoto to Bali: brief overview of RED / REDD negotiations
The UNFCCC, in its principles, declares that “policies and measures (to combat climate change) should … be
comprehensive, cover all relevant sources, sinks and reservoirs of greenhouse gases and adaptation, and
comprise all economic sectors”. The mandate has not been entirely fulfilled by the Kyoto Protocol: indeed, while
it allows developed countries to compensate their emissions with forest management and with forest planting –
both at home and (to a certain extent) through CDM afforestation/reforestation projects in developing countries –
efforts to reduce tropical deforestation were not included. The reasons for the exclusion deforestation-avoidance
projects in developing countries were related to the following issues (Dutschke and Wolf 2007):
• leakage, which refers to indirect effects of the mitigation project on GHG emissions outside the project
(national leakage) or even country boundaries (international leakage), i.e. deforestation may simply be
shifted elsewhere.
• uncertainties of estimates of how much GHG emissions from deforestation has actually been reduced;
• non-permanence, which occurs when carbon “protected” through deforestation avoidance is released to the
atmosphere at a future date, due to natural or anthropogenic disturbance;
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Reducing emissions from deforestation in developing countries:
the new challenge for climate mitigation
• magnitude of possible emission reductions, resulting in industrialized countries to put less effort into
emission reductions from burning of fossil fuels.
However, the situation evolved in recent years: in 2005, at the 11th Meeting of the Parties to the UNFCCC (COP
11, Montreal), Papua New Guinea and Costa Rica, supported by several developing countries, tabled a proposal
for including emissions from avoided deforestation in any kind of compensation scheme under the UNFCCC
(see Fig. 1). As compared to discussion held during Kyoto negotiations, such proposal contained some important
novelties, including the fact that it came from a developing country, that a national approach was envisaged –
thus addressing the national leakage problem – and that methodologies for monitoring deforestation were
improved.
Emissions from deforestation
reference scenario
Reduced
emissions, for
which a “positive
incentive”
(C market? fund?)
may be received
Assessment Period
Reference Period
time
1990?
2005?
Figure 1: simplified representation of how a REDD mechanism could work.
Regarding non-permanence, it can be considered that it becomes a problem only if a country that reduces its
emissions from deforestation is not held liable for later re-emissions by increased deforestation. As for
afforestation and reforestation under the CDM, the solution of temporary crediting could be one solution for
REDD.
Regarding the magnitude of potential reductions, the situation for post-2012 agreements seems different from
that during Kyoto negotiations, as future commitments have not yet been fixed. The influx of REDD credits will
allow to reach more ambitious reductions with less costs, and thus the magnitude of REDD credits should rather
be a hope than a concern (Chomitz 2006). Emissions from deforestation are in the same order of magnitude as all
GHG emissions from the United States. Nobody concerned about climate stability would prefer the US not to
adopt binding commitments, just because this might disrupt the market. Market stability is thus a weak argument
against the inclusion of REDD (Dutschke and Wolf 2007): it must be carefully considered, but it seems solvable.
As a result of the proposal by Papua New Guinea, a two-year process was launched within UNFCCC to consider
ways to reduce emissions from deforestation (RED) in developing countries. During 2006-2007, through two
RED workshops and two submissions of views, Parties started sharing experiences, clarifying the key challenges
in terms of “scientific, technical and methodological issues” and “policy approach and positive incentives” and
identifying useful ways to move forward.
In December 2007, at the Conference of the Parties under the UNFCCC (COP 13, Bali - Indonesia), the RED
issue officially entered into the debate on a global future climate change agreement. The main outcome of this
debate is the “Bali action plan” (http:/unfccc.int/files/meetings/cop_13/application/pdf/cp_bali_action.pdf), i.e.
the beginning of a negotiating process which should end by Dec 2009 in Copenhagen with a global agreement
for the post-2012 period. The pillars of this negotiating process will be: mitigation actions; adaptation actions;
technology development and transfer; financial resources to support mitigation and adaptation actions. In the
context of mitigation, it is recognized the need to incentivise efforts to reduce emissions from deforestation in
developing countries as an important mitigation means in the context of the post-2012 agreement on climate
change. Furthermore, the scope of the mechanism was extended to include also forest degradation (so that
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Giacomo Grassi
“RED” became “REDD”8), and it was noted the need to further explore the role that conservation, sustainable
management of forests and enhancement of forest carbon stocks
Furthermore, a more specific COP decision on “Reducing emissions from deforestation in developing countries:
approaches to stimulate action” (http://unfccc.int/files/meetings/cop_13/application/pdf/cp_redd.pdf) included:
- The setting of an ambitious work programme for 2008 on methodological issues, through new Parties’
submissions and a UNFCCC methodological workshop to be held in Japan (end of June 2008).
- An encouragement for starting “demonstration activities” on REDD, which shall follow an agreed indicative
guidance. The efforts undertaken under these activities will then be considered when addressing further policy
approaches and positive incentives.
The inclusion of tropical deforestation at the heart of negotiation for the future climate change mitigation actions
has a great importance, both scientifically and politically. On the one hand, it is the first attempt within
UNFCCC to address seriously the cause of the 20% of anthropogenic GHG emissions. On the other hand, the
REDD issue will be strategic for promoting an active involvement of developing countries in the future efforts
for mitigating climate change.
Although the issue is extremely complex, and many important political and methodological details still need to
be clarified, all the Parties showed up to now a constructive attitude. The priorities for next years which will
include, among others:
-
developing, also through demonstration activities, innovative ways for addressing the drives of
deforestation and implementing sustainable management systems for tropical forests;
-
make sure that rights of indigenous people are preserved;
-
trying to maximize co-benefits and synergies with other UN conventions (i.e. biodiversity and
desertification);
-
agreeing upon the principles behind the setting of a baseline or reference scenario, against which
emissions of the assessment period will be compared;
-
solving the financial question of the REDD mechanism: market-based, fund-based, or a mix of the two?
-
solving the methodological issues, i.e. develop robust estimates of emissions from deforestation;
-
setting up a credible reporting mechanism, i.e. avoid that “hot air” is generated by the REDD
mechanism.
Given the direct link that the last two priorities have with the scientific community, these issues are addressed in
more detail in the next section.
The challenges for the scientific community
Can we estimate accurately emissions from deforestation?
Emissions from deforestation are calculated as the amount of deforested areas multiplied by the carbon stock
changes per unit of those areas – what IPCC calls “activity data” (AD) and “emission factor” (EF). Recent
literature clearly demonstrates that, through the use of remote sensing techniques, it is possible to assess AD for
tropical deforestation with acceptable confidence since 1990 (e.g., Achard et al 2007), although the cost of data
processing and analysis may be high, depending on the level of accuracy to be reached. More challenging is
estimating forest degradation, since this process is not easily detectable thought remote sensing. In that case
some simplification, as for example considering intact vs non-intact forests (Mollicone et al. 2007a) could help.
On the other hand, data on forest carbon stocks and stock changes (EF) are scarce in the tropics (e.g., Gibbs et al
2007) - generally are available only for broad forest types (IPCC 2006) and in some case are not reliable - thus
significant additional monitoring efforts are necessary. For a more comprehensive analysis, a recent “REDD
sourcebook” (http://www.gofc-gold.uni-jena.de/), prepared by the GOFC-GOLD9 expert group, addresses the
8
While deforestation clearly involves a land use change (from forest to non-forest), forest degradation - in the
UNFCCC context – is generally considered as a loss of C in a forest without involving a land use change.
However, at present it is not fully clear how degradation will be estimated and reported.
9
“Global Observation of Forest and Land Cover Dynamics” (GOFC-GOLD), a technical panel of the Global
Terrestrial Observing System (GTOS), sponsored by FAO, UNESCO, WMO, ICSU and UNEP, is a coordinated
international effort to ensure a continuous program of space-based and in situ forest and land cover observations.
GOFC-GOLD aims at providing a suitable forum to develop consensus among earth observation experts on
92
main methodological issues related to the estimation of GHG emissions from deforestation and forest
degradation in tropical countries.
How to set up a credible reporting system?
Irrespective of the potentialities explored by the scientific literature, the practical feasibility of estimates of
emissions from deforestation at the country level will ultimately depend on the methodological and reporting
requirements of the future REDD mechanism, which in turn will determine the cost of monitoring. Although at
present it is not possible to foresee these requirements in detail, some general features may be easily predicted.
According to the general guidance given for the start of demonstration activities on REDD, estimates of reduced
emissions should be “results based, demonstrable, transparent, and verifiable, and estimated consistently over
time”. To this aim, Parties are encouraged to apply the IPCC’s Good Practice Guidance for Land Use, Land-Use
Change and Forestry (IPCC 2003) as a basis for estimating and monitoring emissions.
This preliminary guidance reflects the fact that, within UNFCCC, when an accounting procedure is foreseen – as
in the Kyoto Protocol and likely also in a future REDD mechanism – sound and robust information must be
provided, typically through a GHG inventory which is subsequently reviewed by independent experts. This
information represents the basis for assessing performance as compared to its commitments or reference
scenario, and therefore represents the basis for assigning eventual incentives or penalties. Thus, it is very likely
that the reporting requirements within a REDD mechanism they will need to fulfill the following current
UNFCCC’s principles for reporting emissions and removals of GHGs:
•
Transparency, i.e. all the assumptions and the methodologies used should be clearly explained and
appropriately documented, so that anybody could verify its correctness.
•
Consistency, i.e. the same methodologies and consistent data sets are used along time.
•
Comparability i.e. estimates of emissions and removals should be comparable among Parties. For this
purpose, Parties should follow the methodologies and standard formats provided by the IPCC.
•
Completeness, i.e. estimates should include – for all the relevant geographical coverage – all the agreed
categories, gases and pools.
•
Accuracy, in the sense that estimates should be systematically neither over nor under the true value, so far as
can be judged, and that uncertainties are reduced so far as is practicable.
Given these requirements, it is important to identify which will be the main challenges for developing countries
when estimating and reporting REDD estimates. Based on the analysis of scientific literature (e.g. Houghton
2005) and of data submitted to UNFCCC (UNFCCC 2005) and to the FAO (FAO 2006), it seems clear that most
developing countries are currently very far from being able to provide complete and accurate estimates of
emissions from deforestation. Although the situation may improve in coming years, it is likely that completeness
and accuracy of REDD estimates will remain a major challenge. The risk is that high uncertainties in input data i.e., area change and C stock change/area - may seriously undermine the credibility of the estimates and therefore
of reduced deforestation as a mitigation option.
In this context, another question arises: how to reconcile the request of robust estimates with the incompleteness
and high uncertainty which will likely characterize the estimates of emissions from deforestation in many
tropical countries?
The conservativeness principle
To address the potential incompleteness the and uncertainties of REDD estimates, and thus to increase their
credibility, it has been proposed to use the principle of conservativeness (e.g., Grassi et al 2007, Mollicone et al
2007b).
This principle is already explicitly used in the context of UNFCCC: for adjustments under Art 5.2 of the
Kyoto Protocol and in the modalities for afforestation and reforestation project activities under the Clean
Development Mechanism (CDM)
The adjustment procedure works as follows (UNFCCC 2006): if an Annex I (industrialized) Party reports to
UNFCCC emissions or removals in a manner that is not consistent with IPCC methodologies and would give
benefit for the Party, e.g. an overestimation of sinks or underestimation of emissions in a given year of the
commitment period, then this would likely trigger an “adjustment”, i.e., a change applied by an expert review
methodological issues related to the use of remote sensing for national-level land cover and land use monitoring,
and related accuracy assessment procedures.
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Giacomo Grassi
team (ERT) to the Party’s reported estimates. In this procedure, conservativeness is ensured by multiplying the
ERT’s calculated estimate (based on, e.g. a default IPCC estimate) by a tabulated category-specific
“conservativeness factors”. Differences in conservativeness factors between categories reflect typical differences
in total uncertainties, and thus conservativeness factors have a higher impact for categories or components that
are expected to be more uncertain (based on the uncertainty ranges of IPCC default values or on expert
judgment). In other words, the conservativeness factor acts to decrease the risk of underestimating emissions or
overestimating removals in the commitment period. In the case of the base year, the opposite applies.
Furthermore, the conservativeness principle is implicitly present also elsewhere. For example, the
Marrakech Accords specify that, under Articles 3.3 and 3.4 of the Kyoto Protocol, Annex I Parties “may choose
not to account for a given pool if transparent and verifiable information is provided that the pool is not a source”,
which means applying conservativeness to an incomplete estimate.
In the REDD context, conservativeness means that – when completeness, accuracy and precision cannot be
achieved – the risk of overestimation of reduced emissions should be avoided or minimized. Although the
usefulness of the conservativeness principle seems largely accepted, its application in the REDD context is still
uncertain. In this context, it was recently proposed a method to apply conservativeness to incomplete and
uncertain REDD estimates (Grassi et al. 2008).
- Addressing incomplete estimates. Achieving the completeness principle will clearly depend on the processes,
pools and gases that need to be reported, and on the forest-related definitions that are applied. However, it is
likely that the most typical and important example of incomplete estimates will arise from the lack of reliable
data for a carbon pool. For example, evidence from official reports (e.g., UNFCCC 2005, FAO 2006) suggests
that only a very small fraction of developing countries currently reports data on soil carbon, even though
emissions from soils following deforestation are likely to be significant in many cases. In practice, if soil is not
accounted for, the total emissions from deforestation will very likely be underestimated in both periods.
However, assuming for the most disaggregated reported level (e.g., a forest type converted to cropland) the same
EF in the two periods, and provided that the area deforested is reduced from the reference to the assessment
period, also the reduced emissions will be underestimated. In other words, although neglecting soil carbon will
cause a REDD estimate which is not complete, this estimate will be conservative (see Fig. 2). However, this
assumption of conservative omission of a pool is not valid anymore if, for a given forest conversion type, the
area deforested is increased from the reference to the assessment period.
200
150
(1) REDD = 75
(estimate complete)
100
(2) REDD = 50
(estimate not complete,
but conservative)
50
Reference
period
Assessment
period
1
0.5
EF Biomass (C stock change/ha)
100
100
EF Soil (C stock change/ha)
50
50
AD: area deforested (M ha)
Emissions (M tC)
emis sions from s oil
emis sions from biomass
0
Figure 2: Exemplification of how neglecting a carbon pool may produce a conservative REDD estimate.
(1) Complete REDD estimate, including the soil pool; (2) incomplete REDD estimate, as the soil pool is not
accounted for. The latter estimate is not complete, but is conservative (from Grassi et al. 2008)
94
- Addressing uncertain estimates. If the Party has implemented standard statistical techniques to quantify
uncertainties (according to the IPCC guidance), these uncertainties may be easily expressed through a confidence
interval. This interval allows assessing the risk of overestimating the “true value”: indeed, in a normal
distribution, if one takes the mean value, there is an equal chance (50%) for over- and under-estimation of the
true value. Then, the confidence interval may be used as a simple way to be conservative, i.e. to decrease the
probability of producing an error in the unwanted direction (i.e. overestimating the emissions in reference period
or underestimating the emissions in the assessment period). For example, by taking the lower bound of the 50%
or 95% confidence interval of emissions in the reference period (i.e., correcting downward the original estimate)
means, respectively, having 25% or 2.5% probability of overestimating the “true” value of the emissions. By
contrast, to be conservative in the assessment period, it should be taken the higher bound of the confidence
interval (i.e., correcting upward the original estimate).
Grassi et al. (2008) presents a more detail description of these approaches, and concludes that addressing
potentially incomplete and highly uncertain REDD estimates through the conservativeness principle has the
following advantages:
-
Increases the scientific robustness, the environmental integrity and the credibility of any REDD mechanism.
By decreasing the risk that economic incentives are given to undemonstrated reductions of emission, the
credibility of any REDD mechanism becomes less constrained by the level of accuracy of the estimates.
This should help convincing policymakers, investors and NGOs in industrialized countries that a robust and
credible reporting of REDD estimates is possible.
-
Rewards the quality of the estimates. Indeed, more accurate/precise estimates of deforestation, or a more
complete coverage of C pool (e.g., including soil), will likely translate in higher REDD estimates, thus
allowing to claim for more incentives. Thus, if a REDD mechanism starts with conservativeness, precision
and accuracy will likely follow.
-
Allows flexible monitoring requirements: since the quality of the estimates is rewarded, it could be
envisaged a system in which - provided that conservativeness is satisfied, - Parties are allowed to choose
themselves what pool to estimate and at which level of accuracy/precision, depending on their own costbenefit analysis and national circumstances.
-
Stimulates a broader participation, i.e. allows developing countries to join the REDD mechanism even if
they cannot provide accurate/precise estimates for all carbon pools or key categories, and thus decreases the
risk of emission displacement from one country to another (international leakeage)
-
Increases the comparability of estimates across countries – a fundamental UNFCCC reporting principle and also the fairness of the distribution of eventual positive incentives.
Conclusions
While recognizing the important role of the current sink of the terrestrial biosphere to “buy time”, at the same
time we should be aware that this is a vulnerable gift. The today’s sink of CO2 is likely to shrink, and for the
terrestrial ecosystems it may even turn into a source under future climate. Reducing the risk of sudden climate
change means reducing emissions significantly from all sectors, including the land use sector. In this context, the
ongoing UNFCCC process on possible ways for reducing emissions from tropical deforestation represents an
extraordinary opportunity for cost-effective mitigation action with relevant potential co-benefits, and therefore
should be prioritized in the future architecture of the post-2012 climate regime. Furthermore, the expected
REDD mechanism potentially represents the most important contribution that forest will provide to the future
climate mitigation actions.
The methodologies for estimating accurately emissions from deforestation exist, although their application in
many developing countries could be severely constrained by the cost. However, when data are incomplete (e.g.
emissions from soil not estimated) or the uncertainties in input data (area change and C stock change/area)
remain high, the conservativeness principle may help to increase the credibility of the resulting estimates.
References
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1990s. Global Biogeochem. Cycles, 18, GB2008, doi:10.1029/2003GB002142.
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deforestation. Environ. Res. Lett. 2 045022
Dutschke M., Wolf R (2007) Reducing Emissions from Deforestation in Developing Countries: the way forward,
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Ebeling J 2006 Tropical deforestation and climate change - Towards an international mitigation strategy Oxford
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Res. Lett. 2 045023
Grassi G 2007 RED estimates: do we need accuracy or conservativeness? Presentation at the SBSTA 26 sideevent:
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Grassi, G., Monni, S., Federici, S., Achard, F. and Mollicone, D. (2008) From uncertain data to credible
numbers: applying the conservativeness principle to REDD. Submitted to Environmental Research Letters.
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Draft 1.3 of background paper for Policy Research Report on Tropical Deforestation. Washington, D.C., World
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Eggleston HS, Buendia L, Miwa K, Ngara T and Tanabe K (eds). IGES, Japan
IPCC-WGI 2007 Climate Change 2007: The Physical Science Basis. Cambridge University Press.
Mollicone D, Achard F, Federici S, Eva H, Grassi G, Belward A, Raes F, Seufert G, Matteucci G and Schulze ED 2007a An incentive accounting mechanism for reducing emissions from conversion of intact and non-intact
forests. Climatic Change 83:477–493
Mollicone D, Freibauer A, Schulze E-D, Braatz S, Grassi G and Federici S 2007b. Elements for the expected
mechanisms on Reduced Emissions from Deforestation and Degradation (REDD) under UNFCCC. Environ.
Res. Lett. 2 045024
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Annex I to the Convention FCCC/SBI/2005/18/Add.2
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FCCC/KP/CMP/2005/8/Add.3 Decision 20/CMP.1
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Decision 2/CP.13.
96
DISTURBANCES IN A CENTRAL EUROPEAN MOUNTAIN FOREST IN THE
CONTEXT OF CARBON DYNAMICS: MORE TEMPORAL DISCONTINUITY
THAN EXPECTED
Georg. GRATZER 1, Splechtna B.E.1,2 and Rudel, B3.
1)
2)
Institute of Forest Ecology, Department of Forest- and Soil Sciences
University of Natural Resources and Applied Life Sciences, Vienna,
Peter-Jordan-Strasse 82, 1190 Vienna, Austria
[email protected]
Current affiliation: Center for Environmental Studies and Nature Conservation
University of Natural Resources and Applied Life Sciences, Vienna,
Peter-Jordan-Strasse 82, 1190 Vienna, Austria
3)
Institute of Surveying, Remote Sensing and Land Information
University of Natural Resources and Applied Life Sciences, Vienna
Peter-Jordan-Strasse 82 1190 Vienna Austria
Georg Gratzer, Splechtna B.E. and Rudel, B
DISTURBANCES IN A CENTRAL EUROPEAN MOUNTAIN FOREST IN THE CONTEXT OF
CARBON DYNAMICS: MORE TEMPORAL DISCONTINUITY THAN EXPECTED
Introduction
Disturbances play a major role in structuring plant communities and in maintaining diversity and
productivity (Pickett and White 1985). Different rates, intensities and sizes of disturbances influence the
resource availability in the disturbed areas and create different opportunities for species as a result of life history
trade-offs (Canham 1989; Denslow 1980; Denslow et al. 1998; Turner et al. 1998; Pearson et a. 2003).
Disturbance return intervals may set temporal limits for community dynamics and may influence diversity and
species coexistence (Connell 1978; Sheil and Burslem 2003; Nakashizuka 2001).
This role of disturbances has long been acknowledged and has led to a number of studies on vegetation
dynamics worldwide. The climatic effect of increased CO2 concentration has prompted a large number of studies
on the global carbon cycle and on the size and dynamics of carbon sinks (e.g. Baker 2007; Malhi et al. 2008).
Only recently, the role of disturbances, both natural and anthropogenic, in the terrestrial carbon cycle has been
identified as a major modifier of global carbon dynamics and sink sizes (Knohl et al. 2002; Kramer et al. 2004;
Magnani et al. 2007; Bond-Lamberty et al. 2007; Kurz et al. 2008). From a carbon sink perspective, disturbances
were grouped into two classes (Knohl et al. 2002): one where carbon is removed fast as in the case of timber
harvesting or fire and another class where the carbon remains on the disturbed site and is slowly decomposed as
in windthrow disturbances. Particularly the latter group has received less scientific attention (but see Kramer et
al. 2004; Knohl et al. 2002). Storms, however, are a critical disturbance agent in the alpine zone of Europe
(Schelhaas et al. 2003). Together with avalanches and biotic disturbance agents like bark-beetles, they account
for most of the disturbances structuring forest communities. While coarse-scale model based assessments of the
role of windthrow disturbances on carbon budgets did not show a strong response (Thürig et al. 2005), more
detailed studies using eddy covariance methods revealed considerable increases of carbon respiration and a net
release of CO2 from windthrows in boreal areas (Knohl et al. 2002). Because of methodological difficulties, sites
disturbed by windthrow have been largely neglected so far, resulting in biased estimates of the sink function of
forests: recently, the average net ecosystem production (NEP) was estimated to be only 56% of peak NEP when
disturbances are accounted for (Magnani et al. 2007). This underlines the important role of disturbances, not only
for the coexistence of species and maintenance of biodiversity but also for carbon sequestration. Two main
questions emerge in this context: (1) what is the effect of different disturbances and different disturbance
intensities on carbon dynamics and (2) what is the spatio-temporal distribution of disturbances? In this paper we
will focus on the second question. This becomes important when the average NEP is to be assessed. At its
simplest, a landscape is in a shifting mosaic steady state and the disturbance regime is static. Assuming such
ergodic conditions, this would allow for a simple estimation of the ratio of disturbed areas and, assuming that the
share of early development phases remains constant over time, a reduction of carbon sequestration. Results of
studies on the disturbance ecology, however, have casted doubt on both assumptions underlying this concept:
firstly, disturbance regimes are not static in time: Westerling et al. (2006) reported a fourfold increase of forest
fires in the Western US since 1986 as a consequence of an altered hydrology. The magnitude of the current
massive mountain pine beetle outbreak in the Western US is partly attributed to climatic change (Kurz et al.
2008). Secondly, owing to their spatial and temporal disturbance characteristics in relation to landscape size and
their recovery intervals, a number of ecosystems are not in steady state conditions (e.g. Turner et al. 1993).
Particularly in Central Europe with its history of intense forest use and exploitation (Glatzel 1999), knowledge
on natural disturbance regimes and its consequences is limited and steady state conditions are assumed to be met
even in small forested landscapes and over short time scales. In this paper we provide a review of recent work
which we conducted in an old-growth mountain forest in Austria (Splechtna et al. 2005, Gratzer et al., in prep.)
to answer the question whether (1) the studied forests show temporally continuous disturbances on decadal and
centennial scales and (2) the spatial scales of the disturbances are small in relation to the landscape. We finally
discuss our results in a landscape equilibrium context.
Methods
Study area
The old growth in Rothwald is the largest and one of the best-documented remnants of never-logged
forest in Central Europe. It is located at the southern slopes of Dürrenstein – a mountain (1878 m asl., 47o 47’ N,
15o 04’ E) in the northern limestone Alps of Lower Austria at an elevation of 900 to 1350 m asl (Figure 1) and
covers 300 ha. The climate is submaritime with long winters and short cool summers. Precipitation pattern is
bimodal with the maximum during the summer months and a second peak in winter. Average annual
precipitation is 2200 mm. Deep and long lasting snow cover leads to a short growing season. The bedrock is
partly comprised of dolomite, partly of banked limestone. Soils were described as a mosaic of Rendzinas and
relictic loams (Rendzic Leptosols and chromic Cambisols) (Zukrigl et al. 1963). Especially in the flatter lowerelevation basin, rich relictic loams are more abundant. The forests of the study area are spruce-fir-beech forests
98
Disturbances in a central European mountain forest in the context of carbon dynamics: more temporal
discontinuity than expected
classified as Helleboro nigri-Fageta or partly as Asperulo -Fageta (Zukrigl 1973; Mucina et al. 1993). The latter
is restricted to parts of the flat basin, where dystric chromic cambisols are more abundant. Typical for the
montane subassociations of both associations are the codominance of silver fir (Abies alba), Norway spruce
(Picea abies) and common beech (Fagus sylvatica). Beech remains the most abundant species, even more so in
the Helleboro nigri-Fagetum, while in the Asperulo-Fagetum conifers have similar standing volume. Conifers
grow 10 to 15 meters taller (up to 58 m, diameters in 1.3 m > 1.5 m), creating the typical two-layered canopy of
these forests. In addition to the three co-dominant tree species, Acer pseudoplatanus, Ulmus glabra, and Taxus
baccata occur sporadically.
99
Vienna
Dürrenstein
~1 ha PSPs
0.25 ha PSPs
N
Plots for gap
dynamics study
0.4
0
0.4
0.8
1.2
1.6 Kilometers
Figure 1: Location of the study area and location of study plots.
100
Material and Methods
The results presented in this paper were derived from two studies: a dendroecological study provided a
disturbance history at centennial scales. At coarser spatial, but shorter temporal scale, we conducted an analysis
of a temporal sequence of aerial photographs in order to characterise the disturbance regime. The methods of the
dendroecological study are described in detail in Splechtna et al. (2005), here, the methods are only briefly
presented.
Dendroecological study
The sampling was carried out on four (100x100m) permanent sample plots (PSPs). Two permanent
sample plots (PSP1 and PSP2) in the lower and flatter, conifer-richer portion of the forest had been established in
1980 and 1943, respectively. The other two plots PSP3 and PSP4 were selected to capture the dynamics of the
beech-dominated stands on slopes (Table 1). Plots were selected to represent a relative uniform site without
strong topographic break lines, or intermeditate-scale edaphic variation. An earlier census of all trees > 1cm in
dbh, (Gratzer & Splechtna unpublished data) revealed that in all plots F. sylvatica dominates by numbers, but to
a lesser extent in terms of basal area. At PSP1 the relative importance of F. sylvatica is lowest, whereas at PSP3
F. sylvatica is clearly dominating. PSP2 and PSP4 have similar species composition, though they represent
contrasting landforms (flat plateau versus steep slope) (Table 1).
Table 1. Characteristics of the four one ha plots (PSPs).
* modified after: Splechtna et al. 2005
PSP1
PSP2
PSP3
PSP4
1050
1020
1090
1200
Aspect (°)
120
110
160
175
Slope (°)
0-10
0-5
15-25
25-30
1- 10 cm dbh
1062
2214
757
1033
≥ 10 cm dbh
197
240
145
206
Total basal area F. sylvatica
22.8
28.9
30.1
27.9
Volume of live F. sylvatica
225
294
346
266
Relative Importance F. Sylvatica
60.0
78.3
93.2
81.4
1 - 10 cm dbh
183
196
21
44
≥ 10 cm dbh
63
31
2
29
Total basal area P. abies
20.2
7.8
0.2
7.4
Volume of live P. abies
210
78
11
42
Relative Importance P. abies
25.3
13.2
1.6
12.1
1 - 10 cm dbh
3
4
0
1
≥ 10 cm dbh
55
30
7
23
Total basal area A. alba
15.0
6.9
3.3
4.5
Volume of live A. alba
154
86
16
38
Relative Importance A. alba
14.8
8.6
5.3
6.6
58
43,6
33,6
39,8
589
470
376
350
Altitude (m)
No. of F. sylvatica
No. of P. abies
No. of A. alba
Total basal area (m²)
Total volume (m³)
We divided all PSPs into one hundred 10 x 10 m squares for systematic sampling. Two cores (from bark
to pith) were extracted from the tree (regardless of species) that was closest to the centre of each quadrat and
greater than 4 cm in diameter. These 100 trees per PSP were later used for the age structure analysis. If a sample
tree was suppressed (directly overtopped by other trees), the next non-suppressed tree was also cored. In total we
cored 118, 115, 106, and 108 cores at the plots PSP1, PSP2, PSP3, and PSP 4, respectively. The 100 nonsuppressed trees per plot were used for the disturbance event analysis. Sample preparation and measurements
were carried out differently for F. sylvatica and for conifers, because the latter were subjected to x-ray
101
densitometry (Schweingruber et al. 1978) so the density data could be used for a study on the climate growth
relationship. Cross-dating was carried out by visual inspection of ring width series using the list method
(Yamaguchi 1991) and statistical evaluation using Cofecha (Holmes 1983).
For defining release events, we used boundary-line release criteria (Black & Abrams 2003, 2004). The
boundary-line method is a running mean method for release detection (Rubino & McCarthy 2004), but differs
from all other available methods by scaling the release relative to the boundary line, i.e. the expected maximum
growth change given the specimens growth immediately preceding the growth change. That way the method
takes into account that the release potential for a specimen is not constant over time, but is much greater when its
growth is strongly suppressed (Black & Abrams 2003).
For calculating the species-specific boundary line, a large data set of tree ring measurements collected
from a variety of sites was needed. Therefore, we supplemented our data with data from other sites in the
limestone Alps in Austria (Grabner et al. 2004) and with all available tree ring series at the International Tree
Ring Data Bank (Contributors of the International Tree Ring Data Bank) for F. sylvatica, P. abies and A. alba.
Prior growth and percent-growth change were calculated for every growth increment, except for the first and last
ten years in each tree ring series due to the constraints of the percent-growth change and prior growth formulas.
We divided the data set into nine prior growth classes (class width 0.5 mm). We averaged the ten highest growth
change values for every growth class, and tested several non-linear functions for their fit to these nine top values.
In the second step, all the releases according to the Nowacki & Abrams criteria were scaled relative to
the expected maximum growth change quantified by the boundary line. Visual inspection of the data showed that
the variation of ring width increased with increasing prior growth, because late-frost events or other short-term
climate extremes or mast years lead to very narrow rings even during periods of high growth (Splechtna &
Gratzer unpublished data). Even after smoothing the growth using 10-year running means (using the Nowacki &
Abrams (1997) formula) this higher year-to-year variation appeared to influence the release detection. Therefore,
we used only growth pulses of at least 50 % growth change according to Nowacki & Abrams (1997) and
accepted only these events as potential releases. Only these potential releases were then scaled relative to the
boundary line. After making these adjustments to the boundary line method, the risk was very low for accepting
climatically induced spurious releases. Therefore, we could select the relatively liberal release criteria for scaling
candidate releases to the boundary line analogous to Black & Abrams (2003). We defined, a moderate release as
any percent-growth change pulse >20 % of the boundary line at the given prior growth rate, and we defined a
major release as any pulse exceeding 50 % of the value of the boundary line. The proportion of cores at coring
height 1.1m, which released every decade was then calculated and displayed as a disturbance chronology
(Lorimer & Frelich 1989).
Gap origin events (Lorimer & Frelich 1989) were defined based on early growth rates of the trees.
When plotted against dbh, there was only a small overlap in recent growth (mean growth of the last 10 years)
between the a-priori field-determined groups of suppressed and dominant trees up to a dbh of 50 cm, indicating
that based on a growth threshold most individuals would be classified correctly. We defined the threshold for
gap origin as the upper 95th percentile of growth rate of suppressed trees over the last ten years (1991:2000),
which corresponded closely with the 5th percentile of dominant trees. For F. sylvatica the threshold was 0.75
mm, and for P. abies and A. alba the threshold was 1.1 mm. Every tree that grew at a higher rate than the
threshold between the years 6 and 15 counted from the pith was considered gap origin.
Analysis of arial photographs
For the analysis of aerial photographs, we used the longest time series with sufficient quality. After
intensive data search and evaluation, photographs from 1962, 1991, and 1996 were analysed. Within the old
growth forest, we selected four plots in order to cover both, the basin and the steeper slope portions of the area
(Fig. 1). The four plots had a total size of 58 ha. An automatically generated model of the canopy surface proved
to be inaccurate and was not able to detect more than 56% of the gaps in a pilot area. Consequently, manual
measurements in a grid of 5 m x 5 m were carried out. The subsequent calculation of the net canopy height for
characterisation of gap structure required an accurate elevation model of the ground surface. We used aerial
photographs taken in winter 1993 and measured a highly accurate digital elevation model using Erdas Imagine
8.5 Stereo Analyst. Measurements for the ground surface model were not carried out in a grid but in a dense net
of well visible ground surface points. In addition to that, surface break-lines were added to the surface model. A
total of 70696 measurements were carried out. For both, the canopy surface models and the ground surface
model, triangulated irregular networks (TIN’s) were calculated using the Terrain Analyst module of
ImageStation (Intergraph 1997b). For each point in time, height differences were calculated from the canopy
height and the ground surface model. The procedure resulted in three grid files (one for each point in time) with
quadrats of 5x5m containing information on canopy height for all four plots.
In the literature, several definitions for canopy gaps were proposed and applied (e.g. Brokaw 1982;
Nakashizuka 1984; Veblen 1985). For this study, we used the 99th percentile of the tree height of all suppressed
trees (Kraft 1884 in Assmann 1961) measured on 10 permanent sample plots in the study area as threshold. A
total of 6820 trees were recorded, the respective height was 14.7 m. Hence we defined gaps as areas with a
canopy surface below 15 m. This is in line with Zukrigl et al. (1963) and Schrempf (1980) who reported that in
102
the study area, trees gain access to the lower layers of the canopy at the same height. The same definition was
also used by Nakashizuka et al. (1995), Tanaka and Nakashizuka (1997) and Fujita et al. (2003a and b).
We derived gap size distributions of the four plots and the three times and tested the distributions for
significant differences between times using a one sample Kolmogorov-Smirnov test. The rates of annual gap
formation and closure were calculated according to Fujita et al. (2003a) by dividing the number of quadrats
which changed their state from canopy to gap or vice versa by the product of the total number of quadrats and
the time interval.
We derived estimates of the fraction of quadrats, which were closed through lateral encroachment and
gap filling from below based on maximum annual height growth rates. We derived these maximum height
growth rates from height index curves (Marschall 1975). For P. abies we calculated maximum annual height
growth rates of 70 cm, for A. alba and F. sylvatica 65 and 55 cm, respectively. When the observed annual
change in canopy height exceeded 70 cm, we assumed lateral encroachment as the cause of canopy closure. For
this analysis we used the period from 1991 – 1996. Within this period, a number of gaps did not show any
closure, neither through lateral encroachment from surrounding canopy trees nor through gap filling from below
through tree regeneration in the gap. In order to show the influence of gap size on lateral encroachment and gap
filling from below, we used gaps showing one or both kinds of gap closure.
Results
The temporal component - disturbance history
The disturbance chronologies including release and gap origin events for the four PSPs revealed a
relatively low proportion of trees showing release in most decades; in few decades more than 20% of trees show
release, and in many decades, fewer than 10% of trees show release (Figure 2). Especially low levels of
disturbance occurred in all four plots between the late 18th century and the early 20th century. However, the exact
timing varied with location. Periods lacking disturbance were longer and more pronounced at stands on flat areas
(PSP1: 1780 – 1910 and PSP2: 1750 – 1890) compared to the stands located on slopes (PSP3: 1840 – 1910 and
PSP4: 1830 – 1940) (Figure 2). At times releases occurred synchronously among all plots in the stand. In
particular, there was an increase in disturbance events during the 20th century, with peaks in three of four PSPs
during the 1930s, 1960s, and 1980s. Also, all plots showed disturbance peaks during the early portion of the
chronology lasting roughly from 1700 to 1760 (Figure 2). Most peaks in releases agreed with the relatively broad
peaks in age distribution (indicated by arrows in Figure 3), but frequent releases during the 1910s and 1930s at
PSP3 were not reflected in the age distribution (Figures 2 and 3).
The temporal pattern of seedling recruitment to the coring height of 1.1 m was discontinuous at all four
plots with periods of no or very little recruitment lasting for as long as several decades (Figure 3). Length and
timing of these periods varied between plots and occurred at PSP1 from 1860 to 1910, at PSP2 from 1740 to
1840, at PSP3 from 1870 to 1950, and at PSP4 from 1810 to 1940.
103
100
40
PSP1
80
60
1980
1960
1940
1920
1900
1880
1860
1840
1820
1800
1780
1760
0
1740
0
1720
20
1700
10
1680
40
1660
20
100
PSP2
80
30
60
1980
1960
1940
1920
1900
1880
1860
1840
1820
1800
1780
1760
0
1740
0
1720
20
1700
10
1680
40
1660
20
100
40
80
PSP3
30
60
1980
1960
1940
1920
1900
1880
1860
1840
1820
1800
1780
0
1760
0
1740
20
1720
10
1700
40
1680
20
Sampling depth (No. of trees)
40
1660
Percentage of trees indicating a disturbance event
30
100
40
PSP4
80
30
60
1980
1960
1940
1920
1900
1880
1860
1840
1820
1800
1780
1760
0
1740
0
1720
20
1700
10
1680
40
1660
20
From: Splechtna et al. 2005.
Figure 2. Proportion of trees with moderate (black) and major (light shaded) release or gap origin events
(dark shaded) relative to the number of trees alive during a given decade for the four one ha plots.
Sampling depth is given for every decade by the lines.
104
1560
1580
1600
1620
1640
1660
1680
1700
1720
1740
1760
1780
1800
1820
1840
1860
1880
1900
1920
1940
1960
1980
2000
14
12
10
8
6
4
2
0
1560
1580
1600
1620
1640
1660
1680
1700
1720
1740
1760
1780
1800
1820
1840
1860
1880
1900
1920
1940
1960
1980
2000
No. of trees
14
12
10
8
6
4
2
0
PSP2
N = 97
PSP3
N = 80
1560
1580
1600
1620
1640
1660
1680
1700
1720
1740
1760
1780
1800
1820
1840
1860
1880
1900
1920
1940
1960
1980
2000
No. of trees
14
12
10
8
6
4
2
0
PSP1
N = 94
PSP4
N = 94
1560
1580
1600
1620
1640
1660
1680
1700
1720
1740
1760
1780
1800
1820
1840
1860
1880
1900
1920
1940
1960
1980
2000
No. of trees
14
12
10
8
6
4
2
0
No. of trees
From: Splechtna et al. 2005.
Figure 3. Number of trees “established” every decade at 1.1m coring height on a regular grid in four 1 ha
plots. N represents the number of sample trees (out of 100) with reliable age estimates kept in the analysis.
Arrows indicate peaks in the corresponding disturbance chronologies.
105
This is supported by the results from the analysis of the aerial photographs: except for one plot, gap area
increased for all plots from 1962 to 1991 and 1996 (Figure 4). Differences in gap area, however, were only
significant on P4 between 1962 and 1991 (two samples Mann-Whitney test, p ≤ 0.05). The increases in gap area
are also reflected in the average canopy heights of the four plots, which decreased from 27.6 ± 7.3 in 1962 to
25.7 ± 8.7 in 1991 and 25.4 ± 9.0 in 1996. P1, P2, and P3 showed significantly different average canopy heights
between 1962 and 1991 but not between 1991 and 1996. In P4, average canopy heights were significantly
different at all three times (p ≤ 0.01). The percentage of gap area in the study plots showed a strong spatial and
temporal variation, ranging from 2.1% (P3, 1962) to 27.1% (P1, 1996); the overall average gap fraction was 12%
(Table 2).
P1
P2
P4
P3
1962
1991
1996
N
200
0
200
400
600
800 Meters
Fig. 4.Gaps (canopy height ≤ 15m, in black) in the four study plots for 1962, 1991 and 1996.
106
Table 2. Characteristics of canopy gaps (canopy height ≤ 15 m) in the four study plots from 1962 until
1996.
P1
P2
P3
1962
1991
1996
1962
1991
1996
1962
7650
9150
9650
1325
3700
1350
3425 10175 12250
850
339
439
95
161
193
88
152
139
243
231
229
5975
5100
6600
225
450
450
375
2125
2125
6025
3575
3325
9
27
22
14
23
7
39
67
88
47
185
209
Gap fraction (%)
21.5
25.7
27.1
5.3
14.9
5.4
2.1
6.1
7.4
3.3
12.3
13.8
-1
2.5
7.6
6.2
5.7
9.3
2.8
2.3
4.0
5.3
1.4
5.3
6.0
Gap area (m²)
Mean area (m²)
Maximum area (m²)
Number of gaps
Gap density (ha )
1991
P4
1996
1962
1991
1996
11425 42825 47875
The spatial component – gap sizes
The frequency distribution of gap sizes is shown in Figure 5. In all plots, small gaps had the highest frequencies.
P1 and P2 showed a bimodal distribution for 1962 and 1991, respectively. The same was found for P4 in the
recordings of 1991 and 1996, although the second peaks were much lower (note the logarithmic scale in Fig. 5).
P1 showed a comparatively low frequency of small gaps in 1962, a second peak in gap size frequencies between
400 and 500 m² and one at 5500 m². The latter peak reflects the large opening probably caused by an avalanche.
100
Plot 1
10
1962
1991
1996
1
log gap density (nr.10 ha -1)
log gap density (nr. 10 ha-1)
100
0.1
1000
log gap area [m²]
1962
1991
1996
1
10000
100
log gap area [m²]
100
Plot 3
1962
1991
1996
10
1
0.1
10
0.1
100
100
Plot 2
1000
Plot 4
1962
1991
1996
10
1
0.1
100
1000
log gap area [m²]
10000
100
1000
log gap area [m²]
10000
Fig. 5. Gap size distributions for four study plots and three observation times (1962, 1991, 1996). Gaps are
defined as canopy heights ≤ 15 m.
107
Discussion
The temporal component
The disturbance history in the studied forests is not static but strongly varies on broad temporal scales.
It is characterized by a relatively low-severity disturbance regime during the 19th century, but higher severity in
the 18th century and a strong increase in the 20th century. The widespread releases in the later 20th century
(1960s and 1980s) coincide with heavy windstorms that hit the area in 1966, 1976, and 1990, indicating that the
disturbance histories inferred from release and gap-origin events are accurate. Disturbance pulses occurred
synchronously on up to three out of four plots distributed across the old growth remnant, indicating that single
events or periods of increased disturbance affect larger areas of the forest. This is also in agreement with earlier
reports that between 1960 and 1980 the standing volume of a portion of 60 hectares of the forest was decreased
by 20 % likely due to windthrow (Schrempf 1985). The synchronous occurrence of release and gap-origin events
also before 1966 indicates additional intense disturbance events affecting larger portions of the forest. Therefore,
we can conclude that disturbances are indeed episodic.
The importance of these episodic disturbances for the dynamics of the forest is illustrated by the
coincidence of recruitment pulses (of seedlings above 1.1 m) with periods of increased disturbance. Our results
suggest that the successful survival and growth to a larger size corresponds well with episodes of increased
disturbance.
The spatial component
The average gap fraction observed in the old-growth forest Rothwald was comparable to those
in temperate forests in North America (Runkle 1990; Spies et al. 1990) and Japan (Yamamoto 1994; Tanaka and
Nakashizuka 1997; Fujita et al. 2003). In Rothwald, the variation of gap characteristics between the plots was
high. This reflects not only differences in size of disturbances but also in disturbance agents: the two largest gaps
in the studied area were caused by avalanches, which occurred before 1962. The gap size distributions in our
study concurred with those reported for boreal (Kneeshaw and Bergeron 1998), temperate (Runkle 1985) and
tropical forests (Brokaw 1985) in featuring many small gaps and few larger openings. The gap size distributions
in our study had a better fit with a power model than a negative exponential model thus confirming the high
frequency of small gaps.
The average gap size was in the upper range of gap sizes found in other temperate forests (McCarthy 2001). The
same refers to the annual gap formation rate in the period of 1991 – 1996 while annual gap formation rates from
1962 – 1991 were much lower. These differences found between the two time periods have to be interpreted with
caution however, since the earlier time span was long enough to allow for some gaps being formed and closed
within the same period. The long interval however, allowed us to obtain information on the canopy structure
prior to the two major storms that occurred in 1966 and 1990, respectively (Splechtna 1994). These two intense
disturbance events, which had disturbed large areas of the surrounding commercially used second growth forests
outside the wilderness area, might explain the strong increase in gap area from 1962 – 1991. Even though no
catastrophic storms hit the area during the second period, gap formation rates were higher than in the first period.
Processes of gap expansion after gaps had been created by the second storm 1990 led to the higher gap formation
during the second period. Through the merging of smaller gaps and through expansion of gaps (see Splechtna et
al. 2005) larger cohorts are formed. Even though this process may cause less increases in ecosystem respiration
than the formation of the same cohort through one larger disturbance, a CO2 release from the gaps can be
assumed (Knohl et al. 2002). The deadwood which remained on the site, however, provides a carbon pool which
buffers the increased respiration, at least to some extent (Davies et al. 2002, Knohl et al. 2002).
The spatio-temporal component
In a simple framework of landscape dynamics, Turner et al. (1993) integrated temporal and spatial parameters of
disturbances. By relating the size of disturbances with the landscape extent and the disturbance interval with the
recovery interval, a state-space diagram of temporal and spatial parameters landscape dynamics was created. In
the following, this framework is applied for the study area and using the data generated by Splechtna et al.
(2005) and Gratzer et al. (in prep.). Using average gap sizes for the disturbance extent, the study area falls well
into the steady state region of the framework (Figure 6). When the merging of gaps and gap extension is
considered, the system shifts into a region with stability but high variance (Figure 6). The size of the landscape
in this example is equal the plot size of one ha. Clearly, this framework is highly scale dependent. For
estimations of the ratio of stand development phases, plots of similar sizes are used (e.g. Zukrigl et al. 1963,
Schrempf 1985). The high temporal variability of disturbances creates a need for including time scales into long
enough to provide reasonable estimates of disturbance intervals for estimations of the effect of disturbance
regimes into assessments of carbon dynamics. Generally, it will be necessary to study the carbon balance and the
subsequent forest development of areas with different disturbance regimes, also including disturbance
interactions between e.g. windthrows and bark beetles. After large disturbances of high magnitude, ecosystems
may experience shifts in their states, e.g. through subsequent soil erosion setting back the system into early
successional stages with long recovery periods. This may result in largely different carbon sequestration and
108
fluxes. Inclusion of successional pathways after disturbances into assessments of carbon dynamics will greatly
improve assessment of sink and source characteristics of forests.
disturbance interval / recovery interval
10.0
5.0
Equilibrium or
steady state
stable, low
variance
Rothwald
1.0
Rothwald
0.5
average gap
size
stable,
very high
variance
merging cohorts
stable high
variance
0.1
stable low
variance
0.05
unstable, bifurcation or crash
0.01
0.25
0.50
0.75
disturbance extent / landscape extent
Figure 6: State space diagram of the spatial and temporal parameters using disturbance characteristics
from Rothwald. Changed from Turner et al. (2003). Disturbance interval was 166 years (derived from
aerial photographs at 1991 and 1996), recovery interval was 227 (the median of canopy accession dates,
Splechtna and Gratzer, unpublished), the disturbance extent was 250 m² (left dot, average gap size in
1996) and 2500 m² (right dot, merging gaps in PSP 2, Splechtna et al. 2005), landscape extent was set to
plot size (1 ha).
109
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111
112
Allegato I
113
114
115
116
117
118
119
120
121
122
123
124
125
126
127
128
129
130
131
132
133
134
Elenco dei Relatori del 44° Corso
Elena Dalla Valle Dip.to Territorio e Sistemi Agro-forestali (TESAF), Facoltà di Agraria,
Università degli Studi di Padova (Italy) e-mail: [email protected]
Davide Pettenella Dip.to Territorio e Sistemi Agro-forestali (TESAF), Facoltà di Agraria,
Università degli Studi di Padova (Italy) e-mail: [email protected]
Domingo Molina
Dep. of Crop and Forest Sciences, University of Lleida (Spain) e-mail:
[email protected]
Ana Isabel Miranda
Dep. de Ambiente e Ordenamento Universidade de Aveiro
(Portugal) e-mail: [email protected]
Gherardo Chirici Dip.to , Facoltà di Agraria, Università degli Studi del Molise (Italy) email: [email protected]
Georg Gratzer
Dip.to , BOKU University, Vien (Austria)
e-mail: [email protected]
Andrea Battisti
Dip.to Agronomia Ambientale e Produzioni Vegetali, Facoltà di
Agraria, Università degli Studi di Padova (Italy) e-mail: [email protected]
Andrea Vannini
Dip.to di Protezione delle Piante, Università degli Studi della Tuscia
(Italy) e-mail: [email protected]
Jean-Francois Boucher
Université du Québec a Chicoutimi - Département des Sciences
fondamentales (Canada) e-mail: [email protected]
Forest Research and Management Institute, Campulung Moldovenesc
Ionel Popa
(Romania) e-mail: [email protected]
Giacomo Grassi
Institute for Environment and Sustainability, Joint Research Centre of
the European Commission, Ispra (Italy) e-mail: [email protected]
135
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Naveh Z. Introduction to landscape ecology as a practical transdisciplinary science of landscape study, planning
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Schaller J. Landscape ecology research and environmental management. Environment and GIS management of a
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Martinkova M.. General aspects of water relations
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Urbinati C., Carrer M. Dendroecologia e analisi della struttura spaziale in una cenosi di "timberline" delle
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ATTI DEL XXXV CORSO - 1998 - La tipologia delle stazioni forestali - Esempio di ecologia applicata
A. Mancabelli, G. Sartori. Roccia madre e suoli del Trentino. Metodologia di rilievo e di studio integrato
dell'ambiente e risvolti tassonomici.
M. S. Calabrese, S. Nardi, Sartori G., D. Pizzeghello, A. Zanella, G. Nicolini. Importanza dell'attività ormonosimile della sostanza umica per una classificazione funzionale degli humus forestali. Applicazione alle
faggete ed abieteti trentini.
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distribuzione delle specie forestali. Applicazioni relative alle indagini delle tipologie forestali.
C. Lasen. Esempi di fitosociologia applicata alla tipologia delle stazioni forestali.
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R. Del Favero. Tipologie forestali: concetti, metodologia e applicazioni. Le esperienze nelle regioni Veneto e
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G. P. Mondino. Presentazione dei tipi forestali del Piemonte.
A. Antonietti. Il metodo fitosociologico applicato alla tipologia delle stazioni forestali in Svizzera.
R.E. Rosselló. Tipi di stazioni forestali in Spagna. Stato dell'arte e prospettive.
M. Bartoli.Confronto tra le tipologie e gli habitat forestali. L'esempio dei Pirenei centrali.
G. Dumè. Il Gruppo di Lavoro sulla tipologia forestale in Francia: risultati e prospettive.
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A. Zanella. Intorno al concetto di "tipologia forestale". Aspettative e realtà.
C. De Siena, M. Tomasi, G. Nicolini. Gli humus forestali del Trentino.
R. Zampedri. Metodologie di interpolazione statistica per una rappresentazione del clima a livello regionale.
ATTI DEL XXXVI CORSO - 1999 - La pianificazione e la gestione del verde urbano
T. Barefoed Randrup. Urban forestry research in Europe.
Z. Borzan, V. Kusan, R. Pernar. Scientific approach to understanding and treatment of amenity trees in urban
forestry.
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P. Raimbault. Assessing and managing urban trees: from scientific concepts to field tecniques.
G. Morelli, G. Poletti. Cenni teorici sulla valutazione della stabilità degli alberi.
L. Benvenuti. Modalità d'intervento e scelta delle tecnologie nella realizzazione delle opere a verde.
ATTI DEL XXXVIII CORSO - 2001 - Monitoraggio ambientale: metodologie ed applicazioni
M. Ferretti. Ecosystem monitoring. From the integration between measurements to the integration between
networks.
A. Benassi, G. Marson, F. Liguori, K. Lorenzet, P. Tieppo. Progetto di riqualificazione e ottimizzazione delle
reti di monitoraggio della qualità dell'aria del Veneto.
S. Fonda-Umani. I sistemi di monitoraggio in aree marine costiere e relative problematiche.
P. Nimis. Il biomonitoraggio della qualità dell'aria tramite licheni.
G. Gerosa, A. Ballarin-Denti. Techniques of ozone monitoring in a mountai forest region: passive and
138
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R. Valentini. Metodologie di studio della produttività primaria di ecosistemi forestali.
C. Urbinati, M. Carrer. L'analisi degli anelli legnosi come strumento per il monitoraggio climatico.
W. Haeberli. Glacier and permafrost monitoring in cold mountain areas as part of global climate related
observation.
R. Caracciolo. Sistema nazionale di monitoraggio e controllo in campo ambientale.
V. Carraro, T. Anfodillo, S. Rossi. I siti sperimentali di "Col de La Roa" e di "Cinque Torri".
ATTI DEL XXXIX CORSO - 2002 - Il fuoco in foresta: ecologia e applicazioni
Giovanni Bovio. La pianificazione antincendi boschivi
Marco Conedera, Marco Moretti, Willy Tinner. Storia ed ecologia degli incendi boschivi al sud delle Alpi della
Svizzera
Thomas W. Swetnam. Fire and climate history in the Western Americas from tree rings
Domingos Xavier Viegas. Fire behaviour models: an overview
Louis Trabaud. Effects of fire on mediterranean plants and ecosystems
Pasi Puttonen. Use of prescribed fire in diversity oriented silviculture
Domingo Molina. Prescribed burning to allow for forest sustainability
Giancarlo Cesti. Tipologie e comportamenti particolari del fuoco: risvolti nelle operazioni di estinzione
Jesús San-Miguel-Ayanz. Methodologies for the evaluation of forest fire risk: from long-term (static) to dynamic
indices
ATTI DEL XL CORSO - 2004 Reti ecologiche: una chiave per la conservazione e la gestione dei paesaggi
frammentati - Ecological networks: a key to the conservation and management of fragmented landscapes
Rob Jongman, The concept of ecological networks: European approaches
Roberto Gambino, Reti ecologiche e territorio
Daniel Franco, Ecological networks: the state of the art from a landscape ecology perspective in the national
framework
Ilse Storch, Wildlife species as indicators: a solution for maintaining "ecological networks" in fragmented
landscapes?
Stefania Zorzi & Silvano Mattedi, Reti ecologiche e fauna selvatica: limiti alla dispersione e loro mitigazione
Duncan McCollin & Janet Jackson, Hedgerows as habitat corridors for forest herbs
Margherita Lucchin, Genetica nelle reti ecologiche: indici e indicatori per la stima della funzionalità
Tommaso Sitzia, La qualità dei corridoi ecologici arborei lineari: indici sintetici di valutazione delle siepi
arboree nel paesaggio agrario
Giuseppe De Togni, Reti ecologiche e pianificazione urbanistica: problemi tecnici e amministrativi
Andrea Fiduccia, Luciano Fonti, Marina Funaro, Lucilia Gregari, Silvia Rapicetta, Stefano Remiero, Strutture di
informazione geospaziale e processi di conoscenza per l’identificazione della connettività ecosistemica
potenziale
Giustino Mezzalira, Progettazione esecutiva e conservazione dei corridoi ecologici arborei
Federico Correale Santacroce, Le reti ecologiche e la Legge Regionale del Veneto 13/2003: linee guida per la
progettazione dei boschi di pianura
ATTI DEL XLI CORSO - 2005 Conoscere il sistema fiume nell’ambiente alpino
Gianfranco Zolin, Corsi d’acqua alpini: ecologia e paesaggio
Giancarlo Dalla Fontana, I processi di formazione del reticolo idrografico
Diego Sonda, Utilizzo di gis per l'analisi del bacino idrografico
Paolo Paiero e Giovanni Papero, La vegetazione rivierasca alpina
Antonio Andrich, Silvia Degli Esposti, Daniele Norbiato, Roberto Dinale,, Marco Borga, Valutazione di alcune
componenti del bilancio idrologico in bacini di tipo alpino
Gian Battista Bischetti, Interazione tra vegetazione e deflusso e stabilità delle sponde
Paolo Billi, I torrenti come condizione di equilibrio morfodinamico e la portata formativa
Vincenzo D'Agostino, Morfologia e dinamica dei corsi d’acqua di montagna
Alessandro Vinello, L’analisi granulometrica dei sedimenti nei corsi d’acqua montani
Lorenzo Marchi, Il trasporto solido di fondo e le colate detritiche: fenomenologia ed effetti sull’assetto dei corsi
d’acqua a forte pendenza
Mario Cerato, Il controllo dei torrenti per mezzo delle opere di sistemazione montana: la ricerca di un
compromesso fra la tutela della naturalità e gli obiettivi di protezione
139
ATTI DEL XLII CORSO - 2006 Stima del carbonio in foresta: metodologie ed aspetti normativi
Davide Pettenella, Giuliana Zanchi,Iinquadramento generale del protocollo di kyoto. Opportunita’ e limiti per il
settore forestale.
Tommaso Anfodillo, Roberto Pilli, Marco Carrer, Vinicio Carraro, Sergio Rossi, Stima della biomassa forestale:
le nuove potenzialita’ delle relazioni allometriche
Giacomo Colle, Flora De Natale, Lucio Di Cosmo, Antonio Floris, Caterina Gagliano, Patrizia Gasparini,
Alessandro Paletto, Gianfranco Scrinzi, Giovanni Tabacchi, Vittorio Tosi, Principali aspetti
metodologici dell’inventario nazionale delle foreste e dei serbatoi forestali di carbonio
Raisa Mäkipää, Integrated method to estimate the carbon budget of forests - nation-wide estimates obtained by
combining forest inventory data with biomass expansion factors, biomass turnover rates and a dynamic
soil c model
Franz Badeck, Forest disturbance factors
Giuliana Zanchi, The article 3.3 and 3.4 activities of the kyoto protocol: requirements and choices
Enrico Pompei, Espansione delle foreste italiane negli ultimi 50 anni: il caso della regione abruzzo
Susanne Klöhn, The role of wood products in the carbon cycle
Mirco Rodeghiero, Metodologie per la stima del carbonio nei suoli forestali
Roberto Pilli, Tommaso Anfodillo, Ilaria Salvadori, Indagine preliminare sullo stock e sulla fissazione del
carbonio nelle foreste del veneto
ATTI DEL XLIII CORSO - 2007 Biomasse forestali ad uso energetico in ambiente alpino: potenzialità e limiti
Mario Lividori, Biomassa legnose- La gestione e gli aspetti selvicolturali.
Eliseo Antonini, Le valutazioni economiche nel settore legno-energia.
Stefano Berti, Le biomassa legnose: il legno.
Lucia Recchia, LCA per filiere legno-energia.
Bernardo Hellrigl, Dendroenergia.
Stefano Grigolato, Pianificazione degli approvvigionamenti in ambiente alpino.
Alexander Eberhardinger, Performance of alternative harvesting methods using feller-buncher system in early
thinnings of Norway Spruce.
Karl Stampfer, Christian Kanzian, Current state and development possibilities of wood chip supply chains in
Ausria.
Walter Haslinger, Combustion technologies and emissions. State-of-the-art small-scale pellets combustion
technologies.
140

disturbi in foresta ed effetti sullo stock di carbonio

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