Urban ecosystem services: Tree diversity and stability of tropospheric ozone removal

Abstract

Urban forests provide important ecosystem services, such as urban air quality improvement by removing pollutants. While robust evidence exists that plant physiology, abundance, and distribution within cities are basic parameters affecting the magnitude and efficiency of air pollution removal, little is known about effects of plant diversity on the stability of this ecosystem service. Here, by means of a spatial analysis integrating system dynamic modeling and geostatistics, we assessed the effects of tree diversity on the removal of tropospheric ozone (O 3) in Rome, Italy, in two years (2003 and 2004) that were very different for climatic conditions and ozone levels. Different tree functional groups showed complementary uptake patterns, related to tree physiology and phenology, maintaining a stable community function across different climatic conditions. Our results, although depending on the city-specific conditions of the studied area, suggest a higher function stability at increasing diversity levels in urban ecosystems. In Rome, such ecosystem services, based on published unitary costs of externalities and of mortality associated with O 3, can be prudently valued to roughly US$2 and $3 million/year, respectively. © 2012 by the Ecological Society of America.

Full text

Ecological Applications, 22(1), 2012, pp. 349–360
Ó2012 by the Ecological Society of America
Urban ecosystem services: tree diversity
and stability of tropospheric ozone removal
FAUSTO MANES,
1,3
GUIDO INCERTI,
1
ELISABETTA SALVATORI,
1
MARCELLO VITALE,
1
CARLO RICOTTA,
1
AND ROBERT COSTANZA
2
1
Department of Environmental Biology, Sapienza University of Rome, Piazzale Aldo Moro, 5, 00185 Rome, Italy
2
Institute for Sustainable Solutions, Portland State University, Portland, Oregon 97207 USA
Abstract. Urban forests provide important ecosystem services, such as urban air quality
improvement by removing pollutants. While robust evidence exists that plant physiology,
abundance, and distribution within cities are basic parameters affecting the magnitude and
efficiency of air pollution removal, little is known about effects of plant diversity on the
stability of this ecosystem service. Here, by means of a spatial analysis integrating system
dynamic modeling and geostatistics, we assessed the effects of tree diversity on the removal of
tropospheric ozone (O
3
) in Rome, Italy, in two years (2003 and 2004) that were very different
for climatic conditions and ozone levels. Different tree functional groups showed
complementary uptake patterns, related to tree physiology and phenology, maintaining a
stable community function across different climatic conditions. Our results, although
depending on the city-specific conditions of the studied area, suggest a higher function
stability at increasing diversity levels in urban ecosystems. In Rome, such ecosystem services,
based on published unitary costs of externalities and of mortality associated with O
3
, can be
prudently valued to roughly US$2 and $3 million/year, respectively.
Key words: air quality; ecophysiology; ecosystem function; GIS; Rome, Italy; sanitary benefits;
tropospheric ozone; urban forest.
INTRODUCTION
Human health and well-being is known to depend on
‘‘ecosystem goods and services’’ (Costanza et al. 1997,
MEA 2005). The concept of ecosystem services was
defined by Daily (1997) as ‘‘the conditions and processes
through which natural ecosystems, and the species that
make them up, sustain and fulfill human life.’’ Since then,
different definitions have been proposed for ‘‘ecosystem
services’’ (Boyd and Banzhaf 2007, Fisher and Turner
2008, Fisher et al. 2009). According to a recent review by
Escobedo et al. (2011), which focused on air pollution
mitigation by urban forests, ecosystem services are
considered the components (including functions) of
urban forests that are directly enjoyed, consumed, or
used to produce specific and measurable human benefits.
Ecosystem services are affected by the relationship
between ecosystem functioning, stability, and biodiversi-
ty (Balvanera et al. 2006, Costanza et al. 2007, Gamfeldt
et al. 2008). The understanding of these relationships is
needed for devising the best management and policy
tools for sustainable use of ecosystem services (Kremen
2005). Urban forests provide important ecosystem
services (Bolund and Hunhammar 1999, Jim and Chen
2008, Young 2010), such as modification of urban
microclimate by lowering temperatures (Pataki et al.
2011a), changing wind patterns, reduction of building
energy use (Akbari 2002), and improvement of local and
regional air quality by removal of atmospheric pollutants
(Nowak et al. 2006). These environmental benefits are
becoming increasingly important, as today more than
half of the world’s population (;3.3 billion people) live
in urban areas. Additionally, according to UN projec-
tions, cities are growing to unprecedented sizes, absorb-
ing nearly all of the growth in the human population over
the next three decades (United Nations Population Fund
2007), with potential implications for biodiversity
conservation issues (Dearborn and Kark 2009).
The role of urban forests in providing ecosystem
services has been investigated in many papers, consid-
ering both basic ecosystem functions, like primary
productivity (Kaye et al. 2006, Pataki et al. 2011b) and
emerging services, such as the improvement of urban air
quality (Yang et al. 2005, Nowak et al. 2006, McDonald
et al. 2007, Escobedo and Nowak 2009). The reduction
of air pollution by urban trees has been recognized as a
cost-effective component of pollution reduction strate-
gies in several urban areas, such as Washington, D.C.,
New York, Baltimore, Atlanta, and Chicago, in the
United States (Nowak et al. 2000, 2006, Yang et al.
2008, Morani et al. 2011), Beijing (Yang et al. 2005),
Santiago de Chile (Escobedo and Nowak 2009), London
(Tiwary et al. 2009), and Toronto (Millward and Sabir
2011). Among the air pollutants removed by urban
forests, tropospheric ozone (O
3
) is dominant in the
Manuscript received 24 March 2011; accepted 5 August
2011; final version received 23 August 2011. Corresponding
Editor (ad hoc): A. Guenther.
3
E-mail: fausto.manes@uniroma1.it
349

photochemical air pollution mixture in urban areas
during summer periods, particularly in Mediterranean
areas (Milla
´n et al. 2000), with negative effects on public
health (Bell et al. 2006, Martuzzi et al. 2006 ). While
robust evidence exists that the physiology of the main
tree functional groups and their abundance and spatial
distribution within cities are the basic parameters
affecting the magnitude and efficiency of O
3
removal
from the urban environment (Escobedo and Nowak
2009), little is known about the effects of tree species
diversity on the magnitude and stability of this
ecosystem service.
The aim of this paper was to quantify and value the
effects of urban tree diversity on the O
3
removal in the
city of Rome (Italy). The underlying hypothesis is that
different tree functional groups exert a complementary
role in stabilizing this emerging ecosystem service over
time, and across different environmental conditions. A
spatial analysis integrating system dynamic modeling
and geostatistics was applied to estimate seasonal and
annual ozone removal by three functional groups of
urban trees, under two climatically different years: the
extremely dry year 2003, and the year 2004, which was
more representative of the average long-term climatic
pattern of the city of Rome (Gerosa et al. 2009).
METHODS
Urban forests in the city of Rome
The city of Rome (418540N, 128290E) extends over an
area of ;1270 km
2
and hosts roughly 2.8 million
inhabitants. Overall, the city is characterized by high
levels of urban traffic and urban expansion, which
largely increased in the last decades. The urban
landscape in Rome is very heterogeneous in terms of
geology, soil, morphology, and land use (Fig. 1a). The
climate is mediterranean, with an average annual
temperature of 15.18C, average annual rainfall of 839
mm, and a typical hot and dry summer period favoring
high tropospheric O
3
concentrations (Manes et al. 2003).
Notwithstanding the long-lasting human impact, span-
ning over .2700 years, and the recent increase of the
urbanized surface, Rome is still considered as one of the
‘‘most green’’ Italian cities, with public green space
covering .20%of the total municipality area and
including a system of nine natural reserves for a total
cover of ;16 000 ha, hosting roughly 1200 plant species
(Celesti-Grapow et al. 2006). Residual fragments of
ancient woodlands still occur within the city boundaries,
hosting a wide set of different tree species ranging from
typical Mediterranean evergreen broadleaf species
(Quercus ilex and Q. suber; hereafter broadleaves) to
deciduous Quercus woods (Q. cerris,Q. frainetto) and
conifer plantations (Pinus pinea). These three groups of
species show important functional differences in their
ecophysiological and phenological traits (Manes et al.
1997, Anselmi et al. 2004), providing an optimal case
study to test the diversity–stability relationship in urban
ecosystems.
As a detailed inventory of the urban forests of the
metropolitan area does not exist, a Landsat 5 TM image
(from 21 July 1999), with a spatial resolution of 30 330
m, was used to assess the distribution of the main tree
functional groups (evergreen broadleaves, deciduous
broadleaves, and conifers) across the city of Rome.
First, a supervised classification of the Landsat image
into 18 land use classes was performed using the TM
bands 3, 4, 5, and 7 by means of a maximum likelihood
algorithm (Anselmi et al. 2003). The overall accuracy of
the classification, calculated through an error matrix,
was 96%. Large green areas can be observed in the
suburban zones, and inside the city center (Fig. 1a). The
most important forested areas, located in the southern
coastal area, are the Castelporziano Presidential Estate,
characterized by high plant community diversity (Seu-
fert et al. 1997), and the Castel Fusano urban park.
Moreover, patches of urban forests are present in the
historical villas, such as Villa Ada, Villa Borghese, and
Veio Park (Celesti-Grapow et al. 2006).
The area covered by each functional group was then
estimated by assigning the urban forest classes in Fig. 1a
to three leaf categories (Fig. 1b). The attribution of the
forest classes to the corresponding functional groups,
being scale dependent, was necessarily affected by some
degree of approximation. Stands characterized by a
dominant species were entirely attributed to the corre-
sponding leaf type: for example, the land use classes
‘‘Holm oak prevailing’’ and ‘‘Cork oak prevailing’’ were
assigned to the ‘‘evergreen broadleaf’’ functional group,
deciduous woods (oak woods dominated by Q. cerris, Q.
frainetto,Tilia cordata,Platanus x acerifolia,and
Robinia pseudoacacia woods) were assigned to the
‘‘deciduous broadleaf’’ group, while Italian stone pine
woods were assigned to the ‘‘conifer’’ group. Deciduous
woods with sclerophyllous species were completely
attributed to the ‘‘deciduous broadleaf’’ category,
because in these stands the evergreen species are mostly
located in the understory layers, thus giving a negligible
contribution to pollutants uptake and deposition
processes (Manes et al. 2007). The areas covered by
mixed conifers and evergreen broadleaved species were
partitioned at 50%between the two leaf categories, while
the attribution of maquis with Holm oak prevailing
areas to ‘‘evergreen broadleaf’’ was limited to 50%of the
total coverage, considering only the area covered by
trees and excluding shrubs and herbaceous species.
As a result, the area covered by the tree functional
groups totaled 7198 ha, corresponding to 5.6%of the
municipality area (Fig. 1b). Deciduous broadleaves
represented the most abundant functional group (3474
ha), followed by evergreen broadleaves (2121 ha) and
conifers (1605 ha). In particular, for the Castelporziano
Presidential Estate, the Castel Fusano urban park, and
Villa Ada, the vegetation cover was: 1080.3 ha, 230.8 ha,
and 38.5 ha for evergreen broadleaves; 2348.9 ha, 215.1
ha, and 35.7 ha for deciduous broadleaves; and 735.2 ha,
429.7 ha, and 42.2 ha for conifers, respectively.
FAUSTO MANES ET AL.350 Ecological Applications
Vol. 22, No. 1

Ozone data
Ozone (O
3
) and nitrogen oxides (NO
x
) concentrations
in air, hourly recorded by 8 and 13 air quality
monitoring stations, respectively, were considered for
the years 2003 and 2004. All categories of air pollution
monitoring sites of the municipal network (urban traffic,
urban background, suburban background, rural back-
ground) were included in the analysis. The spatial
distribution of O
3
concentrations over the survey area
was estimated by applying a spherical co-kriging model
FIG. 1. Urban vegetation of the Municipality of Rome (city limits outlined in black): (a) Landsat 5 TM supervised classification
(modified from Manes et al. [2008]) and (b) distribution and total surface cover of the three tree functional groups (evergreen
broadleaf species, deciduous broadleaf species, and conifers) analyzed in this study within the metropolitan area. Axes report UTM
grid coordinates, WGS84 zone 33 N.
January 2012 351URBAN TREE DIVERSITY AND OZONE REMOVAL

(Isaaks and Srivastava 1989). Daily maps of O
3
concentrations were produced in a Geographic Infor-
mation System (GIS) by using the geostatistical tool
Spatial Analyst in ESRI ArcGIS v. 9.2 (ESRI 2006 ),
where average daily values of O
3
and NO
x
have been
considered as input data. Further, NO
x
was used as an
external drift variable in the co-kriging interpolation
technique.
All maps were produced by smooth interpolation at
30 330 m resolution to be interoperable with the urban
vegetation distribution map (Fig. 1b). In this way, at
each map cell (30 330 m pixel) where the target tree
functional group was located, daily time series of O
3
concentrations were available for analyses. Finally, to
provide a synoptic view of O
3
levels and spatial
variability during 2003 and 2004, annual maps of
‘‘SOMO35’’ (i.e., sum of daily eight-hour running means
of O
3
over 35 ppb; World Health Organization 2008)
were also produced by summing up daily map values.
Modeling tree physiological parameters
The MOCA-Flux (Modeling of Carbon Assessment
and Flux) model, implemented within the object-
oriented software package STELLA II (Costanza and
Gottlieb 1998, Isee Systems 2002), was used to simulate
dynamics of physiological parameters for the three plant
leaf types, for 2003 and 2004. MOCA-Flux is the newly
implemented version of a system dynamic, semi-empir-
ical model, previously applied to simulate functional
responses to changes in air temperature (Vitale et al.
2003) and O
3
stomatal fluxes (Vitale et al. 2005) of Q.
ilex. The MOCA-Flux model is based on the ‘‘big-leaf’’
approach, and it was conceived for estimation of plant
physiological variables including stomatal conduc-
tance (g
s
; mol H
2
Om
2
s
1
), net photosynthesis (P
NET
;
lmol CO
2
m
2
s
1
), leaf transpiration (E;mmol
H
2
Om
2
s
1
), annual net primary productivity (NPP;
g C/m
2
), and leaf area index (LAI; m
2
of leaf/m
2
of
ground). All physiological variables are expressed as a
diurnal average for the photoperiod.
The MOCA-flux model was applied in the current
study due to its demonstrated ability to provide highly
fitting prediction of several physiological parameters
(including stomatal conductance) for different plant
species (Manes et al. 1999, Vitale et al. 2005).
MOCA-Flux calculates stomatal conductance to
water vapor (Eq. 1) by using the Ball et al. (1987 )
algorithm, and corrected by Harley et al. (1992), which
is based on net photosynthesis (P
NET
), relative humidity
(RH), and air carbon dioxide concentration ([CO
2
]
air
),
assumed to be constant at 370 lmol/mol, as follows:
gsðtÞ¼gs0 þmPNETðtÞRHðtÞ
CO2
½
air

ð1Þ
where g
s0
is the minimum stomatal conductance to H
2
O
vapor when P
NET
¼0 and mis an empirical coefficient
that represents the composite sensitivity of conductance
to P
NET
, [CO
2
]
air
, and RH. Net photosynthesis was
calculated as a function of species-specific quantum yield
and solar irradiance by using a semi-empirical model
reported in de Wit et al. (1978; see the Appendix). It is
noteworthy that the leaf area index (LAI) is also related
to net photosynthesis, affecting, in turn, solar irradiance
(see the Appendix). The different modules constituting
the MOCA-Flux model are highly integrated to each
other, thus yielding stable functional interdependences,
minimizing the number of input parameters.
Net primary productivity (NPP) is derived from the
total of diurnal net photosynthesis values integrated in
the phenological time span for each tree species. For
further details on model equations, refer to Vitale et al.
(2003) and (2005). The model has been parameterized
using values of input physiological and structural
variables (see the Appendix) derived from field mea-
surements collected in different sampling sites of the
survey area (Anselmi et al. 2003, Manes et al. 2007,
Vitale et al. 2007), and simulations of daily average g
s
were run for the years 2003 and 2004. The model
validation was based on reference comparison of
simulated O
3
fluxes with eddy covariance measurements,
as reported in Vitale et al. (2005) for summer 2003.
Ozone removal by urban tree functional groups
Stomatal ozone fluxes (FO
3
) were calculated on a
daily time step based on estimated O
3
air concentration
and simulated stomatal conductance to water vapor,
corrected by the diffusibility ratio between O
3
and water
vapor:
FO3ði;pÞ¼gsðiÞ3O3
½
i;p30:613 ð2Þ
where FO
3
is expressed in nmolm
2
s
1
and [O
3
] in parts
per billion (ppb [nmol/mol]), and the indices iand p
refer to the ith day of the reference period and to the pth
location of each tree functional group, respectively.
Stomatal ozone fluxes were referred to unitary area of
soil surface, thus allowing a geographical representation
of the modeling outputs, based on the locations
effectively covered by evergreen broadleaves, deciduous
broadleaves, and conifers within the survey area, as
reported in the vegetation map in Fig. 1b.
The annual time series of FO
3
was integrated over
time at each site to estimate the cumulative amount of
ozone yearly and seasonally taken up in 2003 and 2004
by each tree functional group:
FO3cumðpÞ¼ X
n
i¼1
FO3ði;pÞ3Ph 33600
!
3106ð3Þ
where nis the number of cumulative days, Ph is the
photoperiodinhours,and10
6
is a dimensional
correction factor allowing to express the cumulated
stomatal flux in mmol O
3
m
2
yr
1
,whenFO
3
is
expressed in nmolm
2
s
1
.
To estimate the uncertainty of yearly cumulated
ozone fluxes, the standard deviation of daily stomatal
conductance (i.e., SD
gs
(i) for each ith day of the year)
FAUSTO MANES ET AL.352 Ecological Applications
Vol. 22, No. 1

from six model runs was considered. Consequently,
uncertainty for daily ozone flux was obtained from Eq. 2
for each vegetation type and for each day of the years
2003 and 2004
SDFO3 i;pðÞ¼SDgiðÞ3O3
½
i;p30:613:ð4Þ
Then, uncertainty for yearly ozone fluxes was obtained
by summing up daily contributions.
Significant differences of ozone removal between tree
functional groups were assessed by ANOVA (Duncan
test). Significance was evaluated in all cases at P,0.05.
Average values of the stomatal : total flux ratio for Q.
ilex in the survey area were reported by Gerosa et al.
(2005, 2009), both for 2003 and 2004 (0.29 and 0.43,
respectively). Similar values, ranging from 0.21 and 0.33,
were reported for conifers by Mikkelsen et al. (2004),
though at higher latitude, thus allowing an estimation of
the potential cumulated flux of O
3
removed from
atmosphere by both stomatal uptake and non-stomatal
processes (FO
3t
), at each pth location, as follows:
FO3tðpÞ¼FO3cumðpÞ31
Rð5Þ
where Rvalues were 0.29 and 0.43 for 2003 and 2004,
respectively.
To estimate the total amount of ozone removed from
the atmosphere by evergreen broadleaves, deciduous
broadleaves, and conifers in the Rome municipality, the
cumulated fluxes calculated at each location were
totaled.
Given that all fluxes were referred to 1 m
2
of soil
covered by the target leaf types, and the resolution of the
vegetation map was 30 330 m, then the fluxes in each
pixel were calculated multiplying the flux by the pixel
area, weighted by the relative coverage of each leaf type
in each pixel:
FO3tot ¼X
N
p¼1
FO3tðpÞ3900 ð6Þ
where Nis the number of 30 330 m pixels covered by
the given tree functional type (Fig. 1b), and 900 is the
area in m
2
of each pixel of the map.
The same method was used for estimating the O
3
removal that would have occurred in both years if all the
trees belonged to one single functional group. Three
configurations were considered, in each of which the
total area covered by tree vegetation within the
Municipality of Rome was attributed to one of the
functional groups in Eq. 6.
RESULTS
The years 2003 and 2004 were characterized by
different climatic conditions and ozone pollution levels.
Mean temperatures recorded at the air quality monitor-
ing stations in 2003 were higher than in 2004, for the
months of April (14.0861.88C vs. 12.9861.68C), May
(20.7862.08C vs. 16.1861.88C), June (26.4862.18C vs.
22.0862.28C), July (27.1862.08C vs. 24.1861.98C),
and August (27.8862.18C vs. 24.1861.88C), whereas
total precipitation in 2003 was much lower than in 2004,
especially for April (51 612 mm vs. 104 616 mm), May
(6 63 mm vs. 85 64 mm), June (0 60 mm vs. 22 64
mm), and July (3 61 mm vs. 41 65 mm) (Fig. 2a, b).
Very different spatial (Fig. 3a, b) and temporal (Fig.
2c, d) patterns were observed in ozone concentrations
across the city of Rome during the two years. Mean
monthly values frequently exceeded the threshold of 70
lg/m
3
proposed by the World Health Organization to
quantify O
3
impact on human health (WHO 2008). In
particular, O
3
concentrations were equal to 74 65lg/
m
3
in April, 79 617 lg/m
3
in May, 95 617 lg/m
3
in
June, 97 620 lg/m
3
in July, 94 623 lg/m
3
in August,
and 74 642 lg/m
3
in September, 2003, with corre-
sponding values for 2004 being fairly lower (51 612 lg/
m
3
in April, 67 613 lg/m
3
in May, 74 613 lg/m
3
in
June, 81 615 lg/m
3
in July, 75 613 lg/m
3
in August,
and 58 613 lg/m
3
in September (Fig. 2c, d).
Potential stomatal ozone uptake in 2003 and 2004 was
considerably different among evergreen broadleaves,
deciduous broadleaves, and conifers, showing peculiar
patterns both in time (Fig. 2e–j) and space (Fig. 3c–f ).
In spring of both years, deciduous broadleaves showed
the highest, and conifers showed the lowest, potential
stomatal O
3
fluxes (Fig. 2g, i). In summer 2003,
deciduous broadleaves showed a reduced potential
stomatal O
3
flux due to the reduction of stomatal
conductance under limiting environmental conditions
(Fig. 2g, h), while evergreen broadleaves were able to
maintain high levels of potential stomatal O
3
fluxes (Fig.
2e, f ), and conifers showed an increased O
3
uptake (Fig.
2i, j). In fall, the contribution of the three functional
groups was again different, with higher values estimated
for deciduous broadleaves and lower values for ever-
green broadleaves and conifers (Fig. 2e–j).
The seasonal cumulated stomatal O
3
fluxes (g/m
2
)
maps depicted in Fig. 3c–f, highlighted that evergreen
broadleaves showed cumulated stomatal flux values
fairly constant from spring to summer 2003 (;0.8 g/m
2
),
as slightly increased in 2004 (from 0.4 to 0.5 g/m
2
), as it
can be observed in the Villa Ada urban park and in the
Castelporziano Estate (Fig. 1b).The uptake values of
deciduous broadleaves, shown for example in the large
deciduous forest dominated by Quercus cerrisand
Quercus frainetto, located in the Castelporziano Estate
(southern coastal area; see Fig. 1b), were on average
similar in the spring–summer of 2004 (0.6 g/m
2
) but, in
2003, were higher during spring (0.7 g/m
2
) than during
summer (0.4 g/m
2
). In the same area, conifers showed
ozone uptake values largely increasing from spring to
summer in both years (from 0.2 to 0.8 g/m
2
in 2003, and
from 0.4 to 0.8 g/m
2
in 2004).
Total (stomatal and non-stomatal) ozone uptake by
urban trees was 311.1 Mg in 2003 and 306.9 Mg in 2004,
with an interannual fluctuation between the two years of
January 2012 353URBAN TREE DIVERSITY AND OZONE REMOVAL

,2%(Table 1). However, when the three functional
groups were considered separately, a much higher
variability was found. While annual ozone uptake
decreased of 25%from 2003 to 2004 for evergreen
broadleaves, an opposite pattern was found for decid-
uous broadleaves and conifers. For deciduous broad-
leaves, annual ozone uptake slightly increased by 4.5%
from 2003 to 2004, while, for conifers, the annual ozone
uptake increased considerably by 23%from 2003 to
2004.
The seasonal relative contribution of the three
functional groups to total ozone removal is reported in
FIG. 2. (a, b) Climograms, (c, d) O
3
time series, and (e–j) time series of daily average stomatal conductance ( g
s
; black line) and
potential stomatal O
3
flux (FO
3
; gray line) by the tree functional groups (evergreen broadleaf species, deciduous broadleaf species,
and conifers) in the Municipality of Rome during the years 2003 and 2004.
FAUSTO MANES ET AL.354 Ecological Applications
Vol. 22, No. 1

Table 1, where the removal data were weighted by tree
cover for each group (Mg/ha).
The estimated annual O
3
uptake that would have
occurred in both years if all the trees would belong to the
same functional group is shown in Table 2. Although the
three scenarios are not significantly different from the
actual data set as for the total O
3
removal (2003 þ2004 ),
the inter-year differences (absolute values) in ozone
uptake between 2003 and 2004 for the actual vegetation
pattern (4.2 639.7 Mg) are much lower than for all
single-tree configurations (i.e., 15.1 646.5, 67.0 655.2,
and 90.0 625.4 Mg for the deciduous-only, conifers-
FIG. 3. Spatial patterns of O
3
concentrations and seasonal cumulated stomatal O
3
fluxes for urban trees in the years 2003 and
2004: (a, b) annual SOMO35 (i.e., sum of daily eight-hour running mean of O
3
over 35 ppb; Martuzzi et al. 2006); (c, d) seasonal
cumulated stomatal O
3
uptake by urban trees in spring and summer 2003; and (e, f) seasonal cumulated stomatal O
3
uptake by
urban trees in spring and summer 2004. Axes report UTM grid coordinates, WGS84 zone 33 N.
January 2012 355URBAN TREE DIVERSITY AND OZONE REMOVAL

only, and evergreen-only configurations, respectively;
Table 2).
DISCUSSION
The values of annual ozone uptake normalized by tree
cover estimated for the Municipality of Rome are within
the range reported for other urban areas of different
continents (Nowak et al. 2000, 2006, Yang et al. 2005,
Escobedo and Nowak 2009). The observed differences in
ozone removal ability among the main tree functional
groups (evergreen broadleaves, deciduous broadleaves,
and conifers) can be considered the results of four main
factors: tree cover, plant physiology, leaf season length,
and air ozone concentration. In particular, interannual
and seasonal differences in ozone removal by evergreen
broadleaves are primarily related to the dynamics of
atmospheric ozone concentration, rather than to plant
stomatal conductance, in good agreement with the well-
known drought tolerance of this group (Manes et al.
1997). On the contrary, for deciduous broadleaves,
stomatal conductance was highly affected by the
extreme drought of summer 2003 (Vitale et al. 2007 ),
which produced lower ozone fluxes with respect to
summer 2004. However, in fall, the recovery of stomatal
conductance of this functional group determined higher
ozone fluxes in 2003 than in 2004, in correspondence to
higher ozone levels. For conifers, the slight increase in
ozone uptake in 2004, as compared to 2003, was mainly
related to the winter period. Overall, this group showed
an interannual stability in ozone removal rate, associ-
ated to a rather constant trend of stomatal conductance
(Fig. 2), suggesting a low sensitivity of conifers to the
drought conditions occurring in 2003 (see e.g., Manes
et al. 1997).
Spatial differences in seasonal ozone uptake are
related to the complex interactions between the spatial
distribution of the three functional groups across the
city of Rome and the interannual spatial dynamics of
ozone concentrations in 2003 and 2004. The majority of
ozone removal by urban trees occurred in the southern
coastal area, where the largest urban (Castel Fusano)
and peri-urban (Castelporziano Presidential Estate)
forests are located, and where tree diversity is highest.
However, Fig. 3c–f shows that the urban forest patches
TABLE 1. Seasonal and yearly total ozone removal by the tree functional groups in Rome, Italy, in 2003 and 2004.
Season
and year
Ozone removed, by functional group
Evergreen broadleaves Deciduous broadleaves Conifers Total
Total
(Mg)
Normalized
(Mg/ha)
Total
(Mg)
Normalized
(Mg/ha)
Total
(Mg)
Normalized
(Mg/ha)
Total
(Mg)
Normalized
(Mg/ha)
Winter
2003 9.9 0.0047 0.0 0.0000 2.6 0.0016 12.5 0.0017
2004 11.6 0.0055 0.0 0.0000 8.1 0.0050 19.7 0.0027
Spring
2003 41.9 0.0198 81.9 0.0236 11.4 0.0071 135.2 0.0188
2004 25.7 0.0121 71.4 0.0205 17.2 0.0107 114.2 0.0159
Summer
2003 41.6 0.0196 45.9 0.0132 43.5 0.0271 131.0 0.0182
2004 34.5 0.0163 72.9 0.0210 44.5 0.0277 151.9 0.0211
Fall
2003 10.5 0.0050 16.6 0.0048 5.2 0.0032 32.3 0.0045
2004 5.5 0.0026 7.4 0.0021 8.1 0.0050 21.1 0.0029
All year
2003 103.9 0.0490 144.4 0.0416 62.7 0.0391 311.1 0.0432
2004 77.4 0.0365 151.7 0.0437 77.8 0.0485 306.9 0.0426
Note: Values are expressed as total ozone removal and ozone removal normalized by tree cover.
TABLE 2. Simulation of yearly ozone removal that would have occurred in Rome in 2003 and 2004 if all urban trees belonged to
one single functional group.
Tree functional group
Ozone removal (Mg) Inter-year
difference (absolute
value in Mg)2003 2004 Total
Evergreen broadleaves 352.9
a
613.6 262.8
a
611.8 615.7
a
625.4 90.0
a
625.4
Deciduous broadleaves 299.7
bc
620.1 314.9
b
626.4 614.6
a
646.5 15.1
bc
646.5
Conifers 281.8
c
626.0 348.8
c
629.2 630.6
a
655.2 67.0
ab
655.2
Actual functional groups cover 311.1
b
620.7 306.9
b
619.0 618.1
a
639.7 4.2
c
639.7
Notes: Data are means 6SD, the latter calculated from six simulation runs of the MOCA-Flux (Modeling of Carbon
Assessment and Flux) model. For each column, superscript letters indicate significant differences among tree functional groups
(Duncan test, P,0.05).
FAUSTO MANES ET AL.356 Ecological Applications
Vol. 22, No. 1

in the city center also played an important role by
improving air quality in the most urbanized sites
(Escobedo and Nowak 2009, Morani et al. 2011).
While showing only negligible effects on the total
amount of O
3
removal in 2003–2004, at least for our
specific case study, urban tree diversity significantly
affected the stability of such ecosystem function. Indeed,
a higher interannual difference in ozone uptake was
shown by the simulated single-tree configurations as
compared to the actual tree cover, especially for the
evergreen-only and the conifers-only scenarios. Accord-
ingly, the functional differences (such as response to
drought and length of the leaf-growing seasons) among
tree groups, together with their spatial distribution
across the city of Rome, gave rise to a stable community
function under very different climatic conditions, in
spite of the seasonal fluctuations in ozone uptake of the
different groups.
The observed results fit with a conceptual model of
higher ecosystem function stability at increasing diver-
sity levels. The diversity–stability debate (McCann 2000,
Gamfeldt et al. 2008) is still controversial, with several
long-term experimental studies providing evidence of
both stable and unstable highly diverse ecosystems
(Tilman et al. 2006 ), and only a few papers have
analyzed the effects of biodiversity on urban ecosystem
services (Bolund and Hunhammar 1999, Dearborn and
Kark 2009). In this view, our study is the first reporting
on the diversity–stability debate in urban environments,
with particular focus on the function of air quality
improvement by tree species. The peculiar dynamics of
yearly ozone removal observed for each functional
group (Table 1) shows that under changing climatic
and air pollution conditions, the relative contributions
of these groups to total ozone removal are likely to vary
accordingly. From our analysis, the stabilizing effect of
urban tree diversity on yearly O
3
removal appears as an
emerging property of the urban ecosystem. A common
opinion here is that the diversity–stability relationship is
highly context dependent (Bezener and van der Putten
2007). However, while the impact of urban vegetation
on O
3
removal strictly depends on site-specific condi-
tions, cities usually host high levels of plant diversity
(Celesti-Grapow et al. 2006), which could trigger such
stabilizing effects irrespective of the environmental
characters of the study area. Further studies are required
to address the role of urban tree diversity on the stability
of this ecosystem property in other urban areas, under
different climatic conditions, pollution levels, and urban
tree vegetation pattern.
In general, these results could have important
implications for the development of future management
strategies, such as targeted tree planting in selected
locations or for evaluating the potential benefits to the
stabilizing effect on ozone uptake that could derive from
the replacement of native plant species with ornamental
exotic ones. In particular, the flora of Rome is mainly
composed by native species (Celesti-Grapow and Blasi
1998), and the main exotic tree species are the deciduous
broadleaves Platanus x acerifolia and Robinia pseudaca-
cia, and the confer Pinus pinea, the introduction of
which dates back to the Roman times.
It is worth noting that interannual difference of ozone
removal estimated for the deciduous-only configuration,
although higher, is statistically comparable to the value
obtained for actual vegetation cover. Further, it is
interesting to note the simulations for all single-
functional group configurations, not biased by the
relative tree cover of each functional group (Table 2).
The interannual differences of ozone uptake by the three
functional groups were statistically significant for all
possible pairs, with the exception of the deciduous
broadleaves as compared to conifers (Table 2). At first
glance, such observations might suggest an interchange-
ability of these two tree functional groups with respect
to the stability of the ecosystem function, considering
their overall contributions in the city of Rome.
However, the results of our simulations strongly depend
on the trade-off between the variability of ozone trends
and spatial distribution observed within the city in the
study period. Consequently, it is not possible to state,
for example, that replacing all the conifers with
deciduous broadleaves in a specific urban park, would
not affect the stability of overall ozone removal by
urban trees in the long term. In general, under a
perspective of urban green management, this implies
that the cover of deciduous broadleaves should not be
extended at the expense of the other functional groups.
Further, as the response to drought of the studied
groups represents a key aspect of their functional
complementarity in removing ozone, it can be suggested
that the replacement of drought-tolerant species with
less tolerant ones may negatively affect the stabilizing
effect of diversity.
On the other hand, urban trees may be also associated
to ‘‘ecosystem disservices,’’ such as emissions of volatile
organic compound (Escobedo et al. 2011). Varying
amounts of biogenic volatile organic compounds
(BVOCs) are emitted by different trees (Guenther
1997, Kesselmeier and Staudt 1999), which, in combi-
nation with NO
x
, could have a negative impact on ozone
formation (Owen et al. 2003, Noe et al. 2008). In Rome,
evergreen broadleaves include both strong and medium
monoterpene emitters (like Quercus ilex and Q. suber,
respectively), deciduous broadleaves include both spe-
cies with negligible VOC emissions (Q. cerris) and
medium isoprene emitters (Platanus x acerifolia,Robinia
pseudoacacia), while conifers are dominated by the
medium monoterpene emitter Pinus pinea (Loreto
2002, Loreto et al. 2004, Calfapietra et al. 2009,
Steinbrecher et al. 2009). Different approaches have
been developed for the quantification of BVOCs
emission and reaction potential at urban (Wang et al.
2003, Donovan et al. 2005), regional (Guenther et al.
2000, Leung et al. 2010), and global scales (Guenther
et al. 2006). Nevertheless, evidence on the extent to
January 2012 357URBAN TREE DIVERSITY AND OZONE REMOVAL

which BVOCs may contribute to increase O
3
concen-
trations is still contradictory. Assuming that all BVOCs
emitted by trees react in air to form O
3
, Benjamin and
Winer (1998) estimated the O
3
-forming potential of
urban trees and shrubs. While useful for modelers, this
approach is overly simplified because several com-
pounds in the urban atmosphere can scavenge BVOCs,
thus decreasing the amount of BVOCs available for
ozone formation (Di Carlo et al. 2004). While a
maximum O
3
increment of 10 lg/m
3
due to BVOC
emissions has been estimated for the Mediterranean area
(Thunis and Cuvelier 2000), Nowak et al. (2000)
reported that changing urban vegetation composition
to low BVOC emitters would not produce significant
effects on O
3
concentrations. Furthermore, Calfapietra
et al. (2009) and Fares et al. (2010) have underlined that,
given the scavenging operated by BVOC reactivity with
ozone, trees which emit VOCs may have the capacity to
increase their ozone uptake in comparison with non-
emitting leaves at the same level of stomatal conduc-
tance.
From an economic viewpoint, ecosystem services may
be divided into two general categories: market services
and nonmarket services (Liu et al. 2010). While
measuring market values simply requires monitoring
market data for observable trades, nonmarket services
are much more difficult to value. As we are not aware of
any official data for Europe, the monetary value of the
amount of O
3
removal was estimated using median
externality values for the United States. For ozone, this
value in U.S. dollars per metric ton is O
3
¼$6752/Mg,
and is derived from the median monetized dollar per ton
externality values used in energy decision making from
various studies (Nowak et al. 2006). Based on this value,
for the city of Rome (Table 1), we obtained an estimate
of an annual saving of roughly $2 million.
From a human health viewpoint, a meta-analysis
estimation of the relationship between O
3
concentra-
tions and the percentage of increase in mortality risk
(Bell et al. 2006 ) indicated a 0.3%increase in mortality
per 10 ppb increase in O
3
concentrations. For the
population of Rome, the attributable fraction was
calculated and applied to baseline cardiovascular and
respiratory mortality rates (Martuzzi et al. 2006 ), which
translates to an estimated mortality of roughly 80
deaths/year due to ozone. Using a value of a statistical
life of $1 million results in an annual cost of $80 million.
It has been shown that the urban forests leads to a
decrease in mean O
3
concentrations of ;3%(Alonso
et al. 2011). For the Rome population, this would
correspond to an estimated decrease in mortality of
roughly 3 deaths/year, with a yearly saving of roughly $3
million. These are, of course, rough estimates and are
conservative since they leave out morbidity and other
health effects and use a conservative estimate of the
value of a statistical life (Blomquist 2004 ). Nonetheless,
if the value of ozone uptake by urban trees is calculated
based on the reduction in human health damages, we
obtain a monetary estimate of the same order of
magnitude of that based on externality values.
In conclusion, our results suggest the importance of
urban tree diversity for stabilizing emerging ecosystem
services, such as O
3
removal in urban environment, thus
enhancing human health and well-being. Our findings,
once extended to other urban environments, vegetation
types, and climatic conditions, could bear important
implications for environmental policy and green man-
agement plans oriented at increasing provision of
services in large metropolitan areas, while contributing
to a comprehensive valuing of urban forest diversity.
ACKNOWLEDGMENTS
This research was financed by the European Commission,
FP7-ENV, project HEREPLUS (Health Risk from Environ-
mental Pollution Levels in Urban Systems), Grant Agreement
number 212854. The ArcGIS software package was kindly
supplied by ESRI (Environmental System Research Institute).
The authors thank the Municipality of Rome, Department of
Environmental and Urban Green Policies, for providing the air
quality data and information about the Rome green areas, and
the two anonymous referees for their constructive comments
and suggestions.
LITERATURE CITED
Akbari, H. 2002. Shade trees reduce building energy use and
CO
2
emissions from power plants. Environmental Pollution
116:119–126.
Alonso,R.,M.G.Vivanco,I.Gonza
´lez-Ferna
´ndez, V.
Bermejo, I. Palomino, J. L. Garrido, S. Elvira, S. Salvador,
and B. Artı
´˜
nano. 2011. Modelling the influence of peri-urban
trees in the air quality of Madrid region (Spain). Environ-
mental Pollution 159:2138–2147.
Anselmi, S., M. Chiesi, M. Giannini, F. Manes, and F. Maselli.
2004. Estimation of Mediterranean forest transpiration and
photosynthesis through the use of an ecosystem simulation
model driven by remotely sensed data. Global Ecology and
Biogeography 13:371–380.
Anselmi, S., M. Giannini, A. Allegrini, and F. Manes. 2003.
Nine years’ monitoring of structural and functional vegeta-
tion indices in natural and urban areas. Pages 259–267 in P.
Smits and L. Bruzzone, editors. Analysis of multi-temporal
remote sensing images. Series in remote sensing. Volume 3.
World Scientific Publishing, Toh Tuck Link, Singapore.
Ball, J. T., E. I. Woodrow, and I. A. Berry. 1987. A model
predicting stomatal conductance and its contribution to the
control of photosynthesis under different environmental
conditions. Pages 221–224 in J. Biggins, editor. Progress in
photosynthesis research. Volume 4. Martinus Nijhoff Pub-
lishers, Leiden, The Netherlands.
Balvanera, P., A. B. Pfisterer, N. Buchmann, J.-S. He, T.
Nakashizuka, D. Raffaelli, and B. Schmid. 2006. Quantifying
the evidence for biodiversity effects on ecosystem functioning
and services. Ecology Letters 9:1146–1156.
Bell, M. L., R. D. Peng, and F. Dominici. 2006. The exposure-
response curve for ozone and risk of mortality and the
adequacy of current ozone regulations. Environmental
Health Perspectives 114:532–536.
Benjamin, M. T., and A. M. Winer. 1998. Estimating the ozone-
forming potential of urban trees and shrubs. Atmospheric
Environment 32:53–68.
Bezener, T. M., and W. H. van der Putten. 2007. Ecology:
diversity and stability in plant communities. Nature 446:E6–
E7.
Blomquist, G. C. 2004. Self protection and averting behavior,
values of statistical lives, and benefit cost analysis of
FAUSTO MANES ET AL.358 Ecological Applications
Vol. 22, No. 1

environmental policy. Review of Economics of the House-
hold 2:89–110.
Bolund, P., and S. Hunhammar. 1999. Ecosystem services in
urban areas. Ecological Economics 29:293–301.
Boyd, J., and S. Banzhaf. 2007. What are the ecosystem
services? The need for standardized environmental account-
ing units. Ecological Economics 63:616–626.
Calfapietra, C., S. Fares, and F. Loreto. 2009. Volatile organic
compounds from Italian vegetation and their interaction with
ozone. Environmental Pollution 157:1478–1486.
Celesti-Grapow, L., and C. Blasi. 1998. A comparison of the
urban flora of different phytoclimatic regions in Italy. Global
Ecology and Biogeography Letters 7:367–378.
Celesti-Grapow, L., P. Pyˇ
sek, V. Jaroˇ
sı
´k, and C. Blasi. 2006.
Determinants of native and alien species richness in the urban
flora of Rome. Diversity and Distributions 12:490–501.
Costanza, R., et al. 1997. The value of the world’s ecosystem
services and natural capital. Nature 387:253–260.
Costanza, R., B. Fisher, K. Mulder, S. Liu, and T. Christopher.
2007. Biodiversity and ecosystem services: A multi-scale
empirical study of the relationship between species richness
and net primary production. Ecological Economics 61:478–
491.
Costanza, R., and S. Gottlieb. 1998. Modeling ecological and
economic systems with STELLA: Part II. Ecological
Modelling 112:81–84.
Daily, G. C. 1997. Nature’s services: societal dependence on
natural ecosystems. Island Press, Washington, D.C., USA.
Dearborn, D. C., and S. Kark. 2009. Motivations for
conserving urban biodiversity. Conservation Biology
24:432–440.
de Wit, C. T., J. Goudriaan, H. H. van Laar, F. W. T. Penning
de Vries, R. Rabbinge, H. van Keulen, W. Louwerse, L.
Sibma, and C. de Jonge. 1978. Simulation of assimilation,
respiration and transpiration of crops. Pudoc, Wageningen,
The Netherlands.
Di Carlo, P., et al. 2004. Missing OH reactivity in a forest:
evidence for unknown reactive biogenic VOCs. Science
304:722–725.
Donovan, R. G., H. E. Stewart, S. M. Owen, A. R. MacKenzie,
and C. N. Hewitt. 2005. Development and application of an
urban tree air quality score for photochemical pollution
episodes using the Birmingham, United Kingdom, area as a
case study. Environmental Science and Technology 39:6730–
6738.
Escobedo, F. J., T. Kroeger, and J. E. Wager. 2011. Urban
forest and pollution mitigation. Analyzing ecosystem services
and disservices. Environmental Pollution 159:2078–2087.
Escobedo, F. J., and D. J. Nowak. 2009. Spatial heterogeneity
and air pollution removal by an urban forest. Landscape and
Urban Planning 90:102–110.
ESRI [Environmental System Research Institute]. 2006. Arc-
GIS. Version 9.2. ESRI, California, USA.
Fares, S., M. McKay, R. Holzinger, and A. H. Goldstein. 2010.
Ozone fluxes in a Pinus ponderosa ecosystem are dominated
by non-stomatal processes: Evidence from long-term contin-
uous measurements. Agricultural and Forest Meteorology
150:420– 431.
Fisher, B., and R. K. Turner. 2008. Ecosystem services:
classification for valuation. Biological Conservation
141:1167–1169.
Fisher, B., R. K. Turner, and P. Morling. 2009. Defining and
classifying ecosystem services for decision making. Ecological
Economics 68:643–653.
Gamfeldt, L., H. Hillebrand, and P. R. Jonsson. 2008. Multiple
functions increase the importance of biodiversity for overall
ecosystem functioning. Ecology 89:1223–1231.
Gerosa, G., A. Finco, S. Mereu, M. Vitale, F. Manes, and A.
Ballarin-Denti. 2009. Comparison of seasonal variations of
ozone exposure and fluxes in a Mediterranean Holm oak
forest between the exceptionally dry 2003 and the following
year. Environmental Pollution 157:1737–1744.
Gerosa, G., M. Vitale, A. Finco, F. Manes, A. Ballarin-Denti,
and S. A. Cieslik. 2005. Ozone uptake by an evergreen
Mediterranean forest (Quercus ilex)inItaly.PartI:
Micrometeorological flux measurements and flux partition-
ing. Atmospheric Environment 39:3255–3266.
Guenther, A. 1997. Seasonal and spatial variations in natural
volatile organic compound emissions. Ecological Applica-
tions 7:34– 45.
Guenther, A., C. Geron, T. Pierce, B. Lamb, P. Harley, and R.
Fall. 2000. Natural emissions of non-methane volatile
organic compounds, carbon monoxide, and oxides of
nitrogen from North America. Atmospheric Environment
34:2205–2230.
Guenther, A., T. Karl, P. Harley, C. Wiedinmyer, P. I. Palmer,
and C. Geron. 2006. Estimates of global terrestrial isoprene
emissions using MEGAN (Model of Emissions of Gases and
Aerosols from Nature). Atmospheric Chemistry and Physics
6:3181–3210.
Harley, P. C., R. B. Thomas, J. F. Reynolds, and B. R. Strain.
1992. Modelling photosynthesis of cotton grown in elevated
CO
2
. Plant Cell and Environment 15:271–282.
Isaaks, E. H., and R. M. Srivastava. 1989. An introduction to
applied geostatistics. Oxford University Press, New York,
USA.
Isee Systems. 2002. STELLA II. Isee Systems, Lebanon, New
Hampshire, USA.
Jim, C. Y., and W. Y. Chen. 2008. Assessing the ecosystem
service of air pollutant removal by urban trees in Guangzhou
(China). Journal of Environmental Management 88:665–676.
Kaye, J. P., R. L. McCulley, and I. C. Burke. 2006. Carbon
fluxes, nitrogen cycling, and soil microbial communities in
adjacent urban, native and agricultural ecosystems. Global
Change Biology 11:575–587.
Kesselmeier, J., and M. Staudt. 1999. Biogenic volatile organic
compounds (VOC): an overview on emission, physiology and
ecology. Journal of Atmospheric Chemistry 33:23–88.
Kremen, C. 2005. Managing ecosystem services: what do we
need to know about their ecology? Ecology Letters 8:468–
479.
Leung, D. Y. C., P. Wong, B. K. H. Cheung, and A. Guenther.
2010. Improved land cover and emission factors for modeling
biogenic volatile organic compounds emissions from Hong
Kong. Atmospheric Environment 44:1456–1468.
Liu, S., R. Costanza, A. Troy, J. D’Aagostino, and W. Mates.
2010. Valuing New Jersey’s ecosystem services and natural
capital: a spatially explicit benefit transfer approach.
Environmental Management 45:1271–1285.
Loreto, F. 2002. Distribution of isoprenoid emitters in the
Quercus genus around the world: chemo-taxonomical impli-
cations and evolutionary considerations based on the
ecological function of the trait. Perspectives in Plant Ecology,
Evolution and Systematics 5:185–192.
Loreto, F., P. Pinelli, F. Manes, and H. Kollist. 2004. Impact of
ozone on monoterpene emissions and evidence for an
isoprene-like antioxidant action of monoterpenes emitted
by Quercus ilex leaves. Tree Physiology 24:361–367.
Manes, F., F. De Santis, M. A. Giannini, C. Vazzana, F.
Capogna, and I. Allegrini. 2003. Integrated ambient ozone
evaluation by passive samplers and clover biomonitoring
mini-stations. Science of the Total Environment 308:133–
141.
Manes, F., E. Salvatori, G. La Torre, P. Villari, M. Vitale, D.
Biscontini, and G. Incerti. 2008. Urban green and its relation
with air pollution: ecological studies in the Metropolitan area
of Rome. Italian Journal of Public Health 5:278–283.
Manes, F., G. Seufert, and M. Vitale. 1997. Ecophysiological
studies of Mediterranean plant species at the Castelporziano
Estate. Atmospheric Environment 31:51–60.
January 2012 359URBAN TREE DIVERSITY AND OZONE REMOVAL

Manes, F., G. Seufert, M. Vitale, E. Donato, O. Csiky, and V.
Silli. 1999. Ecophysiological characterization of Citrus
sinensis (L.) Osbeck and relationships with type and amount
of biogenic emissions. Physics and Chemistry of the Earth (B)
24:699–703.
Manes, F., M. Vitale, A. M. Fabi, F. De Santis, and D. Zona.
2007. Estimates of potential ozone stomatal uptake in mature
trees of Quercus ilex in a Mediterranean climate. Environ-
mental and Experimental Botany 59:235–241.
Martuzzi, M., F. Mitis, I. Iavarone, and M. Serinelli. 2006.
Health impact of PM
10
and ozone in 13 Italian cities. World
Health Organization, Regional Office for Europe, Copenha-
gen, Denmark.
McCann, K. S. 2000. The diversity–stability debate. Nature
405:228–233.
McDonald, A. G., W. J. Bealey, D. Fowler, U. Dragosits, U.
Skiba, R. Smith, R. G. Donovan, H. E. Brett, C. N. Hewitt,
and E. Nemitz. 2007. Quantifying the effect of urban tree
planting on concentrations and depositions of PM
10
in two
UK conurbations. Atmospheric Environment 41:8455–8467.
MEA [Millennium Ecosystem Assessment]. 2005. Ecosystems
and human wellbeing: synthesis. World Resources Institute.
Island Press, Washington, D.C., USA.
Mikkelsen, T. N., H. Ro-Poulsen, M. F. Hovmand, N. O.
Jensen, K. Pilegaard, and A. H. Egeløv. 2004. Five-year
measurements of ozone fluxes to a Danish Norway spruce
canopy. Atmospheric Environment 38:2361–2371.
Milla
´n, M. M., E. Mantilla, R. Salvador, A. Carratala
´, and
M. J. Sanz. 2000. Ozone cycles in the Western Mediterranean
Basin: interpretation of monitoring data in complex coastal
terrain. Journal of Applied Meteorology 39:487–508.
Millward, A. A., and S. Sabir. 2011. Benefits of a forested
urban park: What is the value of Allan Gardens to the city of
Toronto, Canada? Landscape and Urban Planning
100(3):177–188.
Morani, A., D. J. Nowak, S. Hirabayashi, and C. Calfapietra.
2011. How to select the best tree planting locations to
enhance air pollution removal in the MillionTreesNYC
initiative. Environmental Pollution 159:1040–1047.
Noe, S. M., J. Pe˜
nuelas, and U
¨. Niinemets. 2008. Monoterpene
emissions from ornamental trees in urban areas: a case study
of Barcelona, Spain. Plant Biology 10:163–169.
Nowak, D. J., K. L. Civerolo, S. T. Rao, G. Sistla, C. J. Luley,
and D. E. Crane. 2000. A modeling study of the impact of
urban trees on ozone. Atmospheric Environment 34:1610–
1613.
Nowak, D. J., D. E. Crane, and J. C. Stevens. 2006. Air
pollution removal by urban trees and shrubs in the United
States. Urban Forestry and Urban Greening 4:115–123.
Owen, S. M., A. R. Mackenzie, H. Stewart, R. Donovan, and
C. N. Hewitt. 2003. Biogenic volatile organic compound
(VOC) emission estimates from a urban tree canopy.
Ecological Applications 13:927–938.
Pataki, D. E., M. M. Carreiro, J. Cherrier, N. E. Grulke, V.
Jennings, S. Pincetl, R. V. Pouyat, T. H. Whitlow, and W. C.
Zipperer. 2011a. Coupling biogeochemical cycles in urban
environments: ecosystem services, green solutions, and
misconceptions. Frontiers in Ecology and the Environment
9:27–36.
Pataki, D. E., H. R. McCarthy, E. Litvak, and S. Pincetl.
2011b. Transpiration of urban forests in the Los Angeles
metropolitan area. Ecological Applications 21:661–677.
Seufert, G., et al. 1997. An overview of the Castelporziano
experiments. Atmospheric Environment 31:5–17.
Steinbrecher, R., G. Smiatek, R. Ko
¨ble, G. Seufert, J. Theloke,
K. Hauff, P. Ciccioli, and R. Vautard. 2009. VOC emissions
from natural and semi-natural vegetation for Europe and
neighbouring countries in the East and South: intra-inter-
annual variability. Atmospheric Environment 43:1380–1391.
Thunis, P., and C. Cuvelier. 2000. Impact of biogenic emissions
on ozone formation in the Mediterranean area: a BEMA
modelling study. Atmospheric Environment 34:467–481.
Tilman, D., P. B. Reicke, and J. M. H. Knopps. 2006.
Biodiversity and Ecosystem stability in a decade-long
grassland experiment. Nature 441:629–632.
Tiwary, A., D. Sinnett, C. Peachey, Z. Chalabi, S. Vardoulakis,
T. Fletcher, G. Leonardi, C. Grundy, A. Azapagic, and T. R.
Hutchings. 2009. An integrated tool to assess the role of new
planting in PM
10
capture and the human health benefits: A
case study in London. Environmental Pollution 157:2645–
2653.
United Nations Population Fund. 2007. State of world
population 2007: unleashing the potential of urban growth.
United Nations Population Fund, New York, New York,
USA.
Vitale, M., S. Anselmi, E. Salvatori, and F. Manes. 2007. New
approaches to study the relationship between stomatal
conductance and environmental factors under Mediterranean
climatic conditions. Atmospheric Environment 41:5385–
5397.
Vitale, M., G. Gerosa, A. Ballarin-Denti, and F. Manes. 2005.
Ozone uptake by an evergreen Mediterranean forest (Quercus
ilex) in Italy. Part II: flux modelling. Upscaling leaf to canopy
ozone uptake by a process-based model. Atmospheric
Environment 39:3267–3278.
Vitale, M., M. Scimone, E. Feoli, and F. Manes. 2003.
Modelling leaf gas exchanges to predict functional trends in
Mediterranean Quercus ilex forest under climatic changes in
temperature. Ecological Modelling 166:123–134.
Wang, Z. H., Y. H. Bai, and S. Y. Zhang. 2003. A biogenic
volatile organic compounds emission inventory for Beijing.
Atmospheric Environment 37:3771–3782.
WHO [World Health Organization]. 2008. Health risks of
ozone form long-range transboundary air pollution. World
Health Organization Regional Office for Europe, Copenha-
gen, Denmark.
Yang, J., J. McBride, J. Zhou, and Z. Sun. 2005. The urban
forest in Beijing and its role in air pollution reduction. Urban
Forestry and Urban Greening 3:65–78.
Yang, J., Q. Yu, and P. Gong. 2008. Quantifying air pollution
removal by green roofs in Chicago. Atmospheric Environ-
ment 42:7266–7273.
Young, R. F. 2010. Managing municipal green space for
ecosystem services. Urban Forestry and Urban Greening
9:313–321.
SUPPLEMENTAL MATERIAL
Appendix
Input values of the MOCA-Flux (Modeling of Carbon Assessment and Flux) model (Ecological Archives A022-022-A1).
FAUSTO MANES ET AL.360 Ecological Applications
Vol. 22, No. 1