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Eprints ID : 17390
To link to this article : DOI: 10.1016/j.ecolind.2016.08.006
To cite this version : Agnan, Yannick and Probst, Anne and Séjalon-
Delmas, Nathalie Evaluation of lichen species resistance to
atmospheric metal pollution by coupling diversity and
bioaccumulation approaches: A new bioindication scale for French
forested areas. (2017) Ecological Indicators, vol. 72. pp. 99-110.
concerning this service should be sent to the repository
Evaluation of lichen species resistance to atmospheric metal pollution
by coupling diversity and bioaccumulation approaches: A new
bioindication scale for French forested areas
Y. Agnana,b,∗, A. Probsta, N. Séjalon-Delmasa,c,∗
aECOLAB, Université de Toulouse, CNRS, INPT, UPS, France
bMilieux Environnementaux, Transferts et Interactions dans les hydrosystèmes et les Sols (METIS), UMR 7619, Sorbonne Universités UPMC-CNRS-EPHE, 4
place Jussieu, F-75252 Paris, France
cLaboratoire de Recherche en Sciences Végétales (LRSV), Université de Toulouse, UPS, CNRS, 31326 Castanet-Tolosan, France
ab s t r a c t
In order to evaluate the metal resistance or sensitivity of lichen species and improve the bioindication
scales, we studied lichens collected in eight plottings in French and Swiss remote forest areas. A total of
92 corticolous species was sampled, grouped in 54 lichen genera and an alga. Various ecological variables
were calculated to characterize the environmental quality – including lichen diversity, lichen abundance,
and Shannon index –, as well as lichen communities. Average ecological features were estimated for each
study site and each of the following variables – light, temperature, continentality, humidity, substrate
pH, and eutrophication – and they corresponded to lichen communities. Based on lichen frequencies, we
calculated the index of atmospheric purity (IAP) and lichen diversity value (LDV). These two bioindication
indices were closely related to lichen diversity and lichen abundance, respectively, due to their calcula-
tion formula. It appeared that LDV, which measures lichen abundance, was a better indicator of metal
pollution than IAP. Coupling lichen diversity and metal bioaccumulation in a canonical correspondence
analysis, we evaluated the resistance/sensitivity to atmospheric metal pollution for the 43 most frequent
lichen species. After validation by eliminating possible inﬂuences of acid and nitrogen pollutions, we
proposed a new scale to distinguish sensitive species (such as Physconia distorta,Pertusaria coccodes,
and Ramalina farinacea) from resistant species (such as Lecanactis subabietina,Pertusaria leioplaca, and
Pertusaria albescens) to metal pollution, adapted to such forested environment.
Atmospheric deposition of chemicals impacts natural ecosys-
tems over a long-term, and biological species are more or less
susceptible to these pollutants (Schulze et al., 1989; Tyler, 1989).
Lichens are considered sensitive organisms because of their biolog-
ical features. The absence of protective cuticle or root system results
in a high sensitivity to anthropogenic disturbances, such as atmo-
spheric pollutants (Bajpai et al., 2010; Conti and Cecchetti, 2001;
Shukla et al., 2014; Szczepaniak and Biziuk, 2003). The loss of lichen
diversity constitutes one of the main markers of atmospheric pol-
lution on the biosphere, as revealed since the ﬁrst observations in
∗Corresponding authors at: ECOLAB, Université de Toulouse, CNRS, INPT, UPS,
E-mail addresses: email@example.com (Y. Agnan),
firstname.lastname@example.org (N. Séjalon-Delmas).
the late 19th century in Paris (Nylander, 1866). Because assessment
of atmospheric pollution is complex and expensive, biomonitoring
is a helpful support technique. Several biomonitoring approaches
are used to evaluate the level of atmospheric pollution, in relation
to lichen diversity (i.e., bioindication; Geiser and Neitlich, 2007;
Pinho et al., 2004) or accumulation of pollutants (i.e., bioaccumu-
lation; Conti et al., 2011; Hissler et al., 2008). Lichens are relatively
good candidates frequently used to monitor atmospheric depo-
sition in various environmental contexts: e.g., forested (Gauslaa,
1995; Giordani et al., 2012), rural (Bosch-Roig et al., 2013; Vonarb
et al., 1990), and urban (Gombert et al., 2004; Loppi et al., 2004)
Atmospheric acid deposition in Europe several decades ago,
linked to man-made SO2and NOx emissions, was responsible
for several disturbances on forest diversity (Schulze et al., 1989).
More speciﬁcally, many authors reported that some lichen species
have disappeared because of their susceptibility to acid pollu-
tants (Piervittori et al., 1997; Sigal and Johnston, 1986). In this
context, a ﬁrst biomonitoring scale was developed in England
and Wales by Hawksworth and Rose (1970), associating common
lichen species for different atmospheric SO2concentrations. More
recently, in Germany, Wirth (1991) developed a toxitolerance index
for more than 750 lichen species also based on the acid pollu-
tion criteria. With the generalized decrease of SO2concentration
in the atmosphere since the 1980′s (Berge et al., 1999), a change
in biomonitoring scale was needed. Several scales were devel-
oped following the relative importance of nitrogen compounds in
the atmosphere (i.e., NOx and NH4;Lallemant et al., 1996; van
Haluwyn and Lerond, 1993). Nevertheless, these various scales
do not take into account other pollutants such as metals (e.g.,
lead, zinc, cadmium) or organic pollutants (e.g., polycyclic aromatic
hydrocarbons [PAH] and polychlorinated biphenyl [PCB]), and lit-
tle is known about the sensitivity or resistance to such pollutants
for lichen species commonly found in northern countries. Conse-
quently, the development of new scales integrating these changes
in sulfur and nitrogen compounds as background levels and the
occurrence of emerging pollutants is therefore required.
In the meantime, several indices of atmospheric air quality were
established based on lichen richness and abundance, such as the
lichen diversity value (LDV; Asta et al., 2002) and the index of
atmospheric purity (IAP; LeBlanc and Sloover, 1970). These indices
attempt to evaluate a general degree of atmospheric pollution. The
limit of such indices, however, is that they do not point to the exact
pollutants caused by disturbance. A qualitative ecological char-
acterization of lichen occurrence should also be employed as an
additional tool to complete the quantitative evaluation, as being
more frequently done.
In this study, we sampled lichen species in open forest sites
from various remote regions of France and neighboring country
to characterize the current degree of recent atmospheric pollution
based on several approaches of lichen biomonitoring. Assuming a
response to a gradient of metal bioaccumulation on lichen richness
and abundance, our main objective was to evaluate the resis-
tance/sensitivity of lichen species to atmospheric metal pollution
by coupling both lichen diversity and bioaccumulation of metals in
a multivariate analysis, and to propose a new resistance/sensitivity
scale adapted to present-day environmental conditions to further
assess the critical loads using lichens.
2. Materials and methods
2.1. Study area
Eight unmanaged open-forested sites were monitored, of which
seven sites from various regions of France, and one site located
in Switzerland (Fig. 1). The French sites (SP 11, EPC 63, EPC 74,
HET 54a, EPC 08, PM 72, and CHS 35) belong to the French moni-
toring network of forest ecosystems RENECOFOR (Réseau National
de suivi des Écosystèmes Forestiers), which is part of the Inter-
national Cooperative Programme Forest network (ICP-Forest). The
sites included both coniferous forests (Abies alba Mill. in SP 11, Picea
abies (L.) H. Karst in EPC 63, EPC 74, and EPC 08, and Pinus pinaster
Aiton in PM 72) and hardwood forests (Quercus petraea (Matt.) Liebl.
in CHS 35 and Fagus sylvatica L. in BEX and HET 54a). Despite the
dominant trees, a mixed of species were found with generally both
coniferous and hardwood trees in each study site.
The sites considered various environmental conditions (Table 1).
The elevation was from 80 m a.s.l for CHS 35–1210 m a.s.l for EPC
74. The Northwestern sites (PM 72 and CHS 35) were inﬂuenced
by an oceanic climate with low annual precipitation (<840 mm),
while the Northeastern (HET 54a and EPC 08) and central (EPC 63)
ones were under semi-continental climate. The climate was more
of mixed inﬂuences for the mountainous sites (SP 11, EPC 74, and
Fig. 1. Location of the study sites sampled for lichen diversity: seven sites are located
in various regions of France and one in nearby Switzerland.
BEX). Several types of bedrock were concerned, from sedimentary
(limestone or sandstone) to magmatic (basalt) substratum.
Metal atmospheric pollution has already been studied for these
sites through surface horizons of soils (Gandois et al., 2010a;
Hernandez et al., 2003), bulk atmospheric deposition (Gandois
et al., 2010b), and lichen bioaccumulation (Agnan et al., 2015).
The metal concentrations registered in lichens collected on the
trees considered for bioindication are given in Table 2. Differences
between sites were observed with a higher anthropogenic inﬂu-
ence in the North-Eastern part of the country, particularly for Pb
and Cd in EPC 08, while a greater dust deposition was observed
in the Southern regions (e.g., in SP 11). The availability of lichen
bioaccumulation data (i.e., metal concentrations in lichens though
accumulation from the environment) was a central part of this
study to determine both lichen resistance and lichen sensitivity
in coupling lichen diversity to the degree of metal concentrations.
2.2. Sampling procedure
Because microclimate and bark properties are known to inﬂu-
ence lichen diversity (Ellis, 2012; Giordani, 2006), each study
site encompassed a representative area of about 250000 m2in
open ﬁeld at the edge of a forest to both maximize the number
of sampling species and preserve the forest inﬂuence (Poliˇ
et al., 2008). Twelve trees avoiding young and disturbed specimens
for lichen sampling (i.e., circumference > 40 cm, inclination < 10 ◦,
trunk without mosses and damages) of various species were sam-
pled (Bargagli and Nimis, 2002; Giordani et al., 2011), including
both deciduous and coniferous trees (Table 1) to improve the repre-
sentativeness of local lichen diversity (Daillant et al., 2007; Deruelle
and Garcia Schaeffer, 1983). We followed the standardized Euro-
pean protocol (EN 16413, 2014), leaving the random sampling to
maximize the number of lichen species by increasing the tree diver-
sity (Moreau et al., 2002). Since we aimed to evaluate the metal
resistance and sensitivity of lichens by combining bioaccumulation
and diversity approaches, we thus followed the same procedure
as for bioaccumulation study (Agnan et al., 2015). The four cardi-
nal points of the tree trunks were sampled using a ladder grid of
ﬁve vertical squares of 10 cm ×10 cm to cover an area of 500 cm2
per tree side and a total area of 24000 cm2(i.e., 240 squares) for
Summary of geographical and environmental characteristics for each study sites.
site coordinates elevation (m) annual
lithology tree species sampled
SP 11 2◦05′40′’E/ 42◦52′15′’N 990 1200 limestome/marble Abies alba Mill., Corylus avellana L.,
Fagus sylvatica L., Fraxinus excelsior L.,
Malus pumila Mill.
EPC 63 2◦58′05′’E/ 45◦45′00′’N 950 1100 basalt Crataegus monogyna Jacq., Fraxinus excelsior L.,
Picea abies (L.) Karst., Pinus sp.
EPC 74 6◦21′00′’E/ 46◦13′30′’N 1210 1300 sandstone/schist Abies alba Mill., Acer sp., Fagus sylvatica L.,
Picea abies (L.) Karst.,
Prunus avium L., Salix sp., Sorbus aucuparia L.
BEX 6◦58′30′’E/ 46◦13′00′’N 945 1000 limestone/schist Acer sp., Betula pendula Roth, Fagus sylvatica L.,
Fraxinus excelsior L., Salix sp.
HET 54a 6◦43′10′’E/ 48◦30′50′’N 320 900 limestone Fagus sylvatica L., Fraxinus excelsior L., Quercus sp.
EPC 08 4◦47′50′’E/ 49◦57′00′’N 475 1300 clay loam Betula pendula Roth, Corylus avellana L.,
Fagus sylvatica L., Picea abies (L.) Karst.,
Prunus avium L., Quercus sp., Rhus hirta (L.) Sudw.,
Salix caprea L., Syringa vulgaris L.
PM 72 0◦20′00′’E/ 47◦44′25′’N 155 800 schist Castanea sativa Mill., Pinus pinaster Ait.,
Quercus petraea (Mattus.) Liebl., Quercus rubra L.
CHS 35 1◦32′50′’W/48◦10′10′’N 80 840 clay Fagus sylvatica L., Pinus pinaster Ait.,
Quercus petraea (Mattus.) Liebl.
Summary of metal bioaccumulation (mean ±standard deviation, in mg g−1) in three foliose lichen species (i.e., X. parietina,P. sulcata, and H. physodes) from the investigated
forest areas (from Agnan et al., 2015).
element SP 11 EPC 63 EPC 74 BEX HET 54a EPC 08 PM 72 CHS 35
Al 2364.1 ±1054.3 988.3 ±299.4 1126.7 ±412.0 1157.9 ±277.8 192.4 ±603.1 1072.7 ±591.1 426.1 ±46.0 397.7 ±62.2
As 0.7 ±0.2 0.7 ±0.4 0.3 ±0.1 0.3 ±0.0 0.3 ±0.1 0.6 ±0.2 0.2 ±0.0 0.2 ±0.0
Cd 0.1 ±0.0 0.1 ±0.0 0.7 ±0.7 0.1 ±0.1 0.2 ±0.2 0.6 ±0.1 0.4 ±0.1 0.1 ±0.1
Co 0.4 ±0.2 0.3 ±0.1 0.4 ±0.1 0.3 ±0.1 0.3 ±0.1 0.4 ±0.1 0.1 ±0.0 0.2 ±0.0
Cr 3.7 ±1.3 2.2 ±1.0 1.8 ±0.6 2.8 ±1.0 1.8 ±0.7 2.5 ±0.9 0.8 ±0.1 0.7 ±0.1
Cs 0.3 ±0.1 0.3 ±0.2 0.2 ±0.1 0.2 ±0.1 0.2 ±0.1 0.2 ±0.1 0.1 ±0.0 0.1 ±0.0
Cu 4.7 ±0.9 6.9 ±2.1 10.4 ±3.2 7.1 ±2.4 7.9 ±3.1 7.3 ±1.0 7.2 ±1.0 5.1 ±1.5
Fe 1347.1 ±595.6 759.1 ±262.2 618.5 ±208.4 687.0 ±158.2 617.6 ±334.8 631.0 ±328.5 278.6 ±35.7 240.9 ±36.4
Mn 29.3 ±14.6 25.1 ±4.0 142.0 ±126.8 102.7 ±74.7 69.5 ±73.2 45.7 ±11.3 45.3 ±4.4 346.8 ±117.7
Ni 1.7 ±0.5 1.3 ±0.5 2.1 ±0.7 2.1 ±0.7 1.7 ±0.6 2.0 ±0.3 0.6 ±0.1 1.4 ±0.3
Pb 2.3 ±1.5 2.5 ±1.1 7.3 ±3.5 5.1 ±2.7 16.1 ±13.4 5.2 ±1.1 1.5 ±0.3 6.2 ±10.4
Sb 0.1 ±0.1 0.1 ±0.0 0.1 ±0.0 0.2 ±0.0 0.2 ±0.1 0.3 ±0.1 0.2 ±0.0 0.1 ±0.0
Sn 0.4 ±0.2 0.3 ±0.2 0.5 ±0.1 0.6 ±0.2 0.4 ±0.2 0.7 ±0.1 0.3 ±0.0 0.2 ±0.0
Sr 8.5 ±3.4 44.4 ±29.8 16.9 ±9.3 16.8 ±5.6 10.4 ±6.3 10.7 ±2.0 4.4 ±0.4 30.7 ±22.5
Ti 187.9 ±82.9 123.2 ±51.2 62.6 ±20.5 80.3 ±19.8 85.6 ±43.8 73.3 ±42.5 32.3 ±2.6 33.6 ±5.4
V 4.1 ±2.0 2.4 ±0.5 2.4 ±0.7 2.2 ±0.5 2.6 ±1.3 2.4 ±0.4 1.0 ±0.1 1.3 ±0.2
Zn 22.1 ±9.6 30.0 ±18.1 69.0 ±43.4 35.1 ±14.4 47.9 ±20.8 108.4 ±11.8 72.8 ±16.2 30.0 ±4.2
each study site (Asta et al., 2002;Fig. 2). The ladder was placed
at minimum 1 m above the ground level to avoid soil inﬂuence
(Bargagli and Nimis, 2002). We determined the presence of lichen
species in each 100 cm2noticed in a sampling sheet: 0 if absent, 1
if present. This allowed obtaining the frequency of each species by
site, averaging all values: from 0 (totally absent in the study site)
to 1 (present in every 10 cm ×10 cm squares). The average values
are given in Table 3. We used a 10- or 30-fold hand lens to identify
all the species. Lichen specimens were collected using a knife, and
preserved in a plastic bag until complete identiﬁcation.
2.3. Species identiﬁcation
Lichen species identiﬁcation was performed in laboratory using
a stereomicroscope (from 20- to 60-fold) and microscope (100-
fold). Determination guides (Clauzade and Roux, 1985; Dobson,
2011; Smith et al., 2009; van Haluwyn and Lerond, 1993), and
chemicals – potassium hydroxide 10% (K), sodium hypochlorite
(C), and paraphenylenediamine (P) – were used to distinguish the
different genera and/or species. Only genera were identiﬁed for
immature specimens. Conversely, we identiﬁed the sub-species
when possible. The nomenclature used was based on Roux (2012).
2.4. Index calculations and statistical treatment
For each site, we determined the number of species found and
the abundance of each species calculated by adding each frequency,
determined using the ﬁeld ladder grid (see above). We also cal-
culated the Shannon’s diversity index H’ based on the following
H’ = −
where piis the proportion of characters of the species i, and R is the
Two bioindication indices were calculated: the lichen diversity
value (LDV; Asta et al., 2002), which represents the sum of frequen-
cies, and the index of atmospheric purity (IAP; LeBlanc and Sloover,
1970) as follows:
where n is the number of species, Qiis the ecological index of each
species i (corresponding to the total number of companion species
present at all studied sites), and fiis the frequency of species i.
Site and average (avg.) frequencies for each lichen species.
species code SP 11 EPC 63 EPC 74 BEX HET 54a EPC 08 PM 72 CHS 35 avg.
(Ach.) A. Massal.
Age 0.092 0.071 0.067 0.021 0.031
(Pers.) Ertz et Tehler
Ava 0.025 0.003
(Hoffm.) Coppins et Scheid.
Apu 0.200 0.008 0.242 0.058 0.021 0.066
(Borrer) R. C. Harris
Abi 0.054 0.007
(Pers.) A. Schneid.
Aat 0.063 0.008
Ara 0.171 0.021 0.054 0.021 0.033
(A. Massal.) Anzi
Aco 0.013 0.002
Bdi 0.104 0.046 0.019
Csa 0.029 0.046 0.009
(Ehrh. ex Hedw.) Th. Fr.
Cce 0.004 0.001
(Hudson) Th. Fr.
Cfe 0.008 0.029 0.005
Cco 0.154 0.019
Cre 0.025 0.003
(Hoffm.) Müll. Arg.
Cvi 0.046 0.006
(Turner ex Sm.) Mig.
Chf 0.075 0.009
(L.) J. R. Laundon
Cca 0.117 0.242 0.042 0.133 0.029 0.063 0.067 0.086
Cﬁ 0.050 0.317 0.171 0.046 0.073
(Turner et Borrer ex Sm.) Ertz
Dde 0.117 0.025 0.046 0.023
Ecr 0.196 0.024
Epr 0.025 0.238 0.104 0.108 0.058 0.042 0.072
Fuscidea cyathoides subsp. corticola
(Fr.) Cl. Roux comb. nov.
Fcy 0.033 0.004
(Borrer ex Sm.) Ach.
Gel 0.175 0.022
Gsc 0.042 0.242 0.042 0.041
(Neck.) J. R. Laundon
Hoc 0.042 0.005
(Ach.) M. Choisy
Hsc 0.013 0.002
Hph 0.067 0.358 0.313 0.017 0.075 0.104
Hla 0.008 0.001
Coppins et P. James
Lsu 0.008 0.038 0.050 0.012
Lab 0.067 0.008
Lal 0.054 0.033 0.011
Lar 0.113 0.254 0.017 0.033 0.052
Aptroot et Herk
Lba 0.063 0.038 0.013
Lca 0.088 0.004 0.008 0.013
Lch 0.150 0.008 0.246 0.788 0.054 0.071 0.004 0.165
van Herk et Aptroot
Lcm 0.133 0.017
Nyl. ex Cromb.
Lcn 0.025 0.004 0.021 0.006
Table 3 (Continued)
species code SP 11 EPC 63 EPC 74 BEX HET 54a EPC 08 PM 72 CHS 35 avg.
Ldi 0.025 0.003
Lex 0.008 0.033 0.005
Lha 0.042 0.005
Lho 0.021 0.003
Lit 0.075 0.009
Lle 0.008 0.001
Lsc 0.025 0.003
Lsr 0.025 0.003
Lecidea sp. Lec 0.021 0.003
(Ach.) M. Choisy
Lel 0.013 0.692 0.038 0.113 0.107
Lic 0.450 0.238 0.483 0.142 0.679 0.458 0.625 0.671 0.468
Lte 0.163 0.020
(Lamy) Sandler et Arup
Mgl 0.121 0.179 0.042 0.442 0.121 0.142 0.174
(DeNot.) O. Blanco, A. Crespo, Divakar,
Essl., D. Hawksw. et Lumbsch
Mea 0.008 0.001
(Nyl.) O. Blanco, A. Crespo, Divakar,
Essl., D. Hawksw. et Lumbsch
Meu 0.138 0.008 0.018
(Flagey ex H. Olivier) O.Blanco, A.
Crespo, Divakar, Essl., D. Hawksw. et
Mla 0.004 0.001
Mpr 0.021 0.003
(Pers.) R. C. Harris
Npu 0.013 0.013
Oan 0.013 0.021 0.004
(L.) A. Massal.
Opa 0.025 0.003
Ochrolechia pallescens subsp. parella
Opp 0.004 0.029 0.004
Och 0.033 0.029 0.008
Ochrolechia sp. Otu 0.029 0.004
Oru 0.038 0.005
Psl 0.075 0.442 0.213 0.579 0.488 0.454 0.050 0.008 0.289
(Taylor) Poelt et Vˇ
Pca 0.096 0.113 0.026
Pab 0.004 0.001
(Huds.) M. Choisy et Werner
Pal 0.033 0.175 0.021 0.083 0.039
Paa 0.125 0.046 0.058 0.029
Pco 0.008 0.488 0.071 0.071
(DC.) J. R. Laundon
Pﬂ 0.013 0.002
Phe 0.013 0.002
Pli 0.013 0.025 0.005
Ppe 0.088 0.011
(Leight.) B. de Lesd.
Psm 0.154 0.019
Par 0.004 0.117 0.221 0.046 0.048
(Fr.) H. Olivier
Pad 0.025 0.417 0.192 0.154 0.098
Table 3 (Continued)
species code SP 11 EPC 63 EPC 74 BEX HET 54a EPC 08 PM 72 CHS 35 avg.
Pcl 0.267 0.033
Plp 0.004 0.001
Pte 0.013 0.229 0.030
(With.) J. R. Laundon
Phy 0.025 0.008 0.004
Pdi 0.042 0.005
Physconia sp. Pen 0.013 0.002
(Neck.) Elix et Lumbsch
Pac 0.025 0.196 0.028
Pfu 0.063 0.304 0.046
Psb 0.021 0.003
Pla 0.083 0.010
Rfr 0.175 0.329 0.017 0.033 0.069
Rfs 0.004 0.001
(Hue) J. R. Laundon
Scr 0.117 0.075 0.024
Tcr 0.046 0.006
Tat 0.021 0.003
Usnea sp. Usn 0.004 0.046 0.006
(L.) Th. Fr.
Xpa 0.013 0.025 0.017 0.075 0.025 0.019
(Pers. ex Ach.) Poetsch et Schied.
Zvi 0.071 0.009
Pvi 0.188 0.092 0.146 0.304 0.188 0.063 0.017 0.124
all 2.854 3.313 3.596 3.200 3.767 2.171 2.067 1.867
Fig. 2. Sampling procedure using a 10 cm ×50 cm grid on the tree trunk in the four
A Student t-test was applied on lichen diversity between each
tree genus (a= 0.05). The lichen frequencies did not follow a normal
distribution (Shapiro-Wilk test); then, data were log-transformed
for the multivariate analyses. Principal component analysis (PCA)
was performed on ecological and environmental data (Dobson,
2011; Nimis and Martellos, 2008; Smith et al., 2009; Wirth, 2010)
based on lichen species frequency. Canonical correspondence anal-
ysis (CCA) was used to evaluate the resistance or sensitivity of the
43 most abundant lichen species to metal atmospheric pollution
based on species frequency. Statistical analyses were carried out
using RStudio 0.98 (RStudio Inc., Boston, Massachusetts, USA) and
ade4 package (Dray and Dufour, 2007).
3.1. Ecological indices
3.1.1. Lichen and tree diversities
The identiﬁed lichen species and their respective frequency for
each study site are reported in Table 3. A total of 54 lichen gen-
era, distributed in 92 corticolous species, and an alga (Pleurococcus
viridis Ag.) were sampled (Fig. 3a). The most abundant species
were Lepraria incana (L.) Ach. (observed in 8 sites with a total fre-
quency of 3.75), Parmelia sulcata Taylor (8 sites, frequency of 2.31),
Lecanora chlarotera Nyl. (7 sites, frequency of 1.32), and Melanelixia
glabratula (Lamy) Sandler & Arup (6 sites, frequency of 1.05). Some
species were found in only one site with a very low frequency
(<0.005): e.g., Ramalina fastigiata (Pers.) Ach., Physcia leptalea (Ach.)
DC, and Parmeliopsis ambigua (Wulfen) Nyl. Overall, the lichen
species were distributed into 64 crustose, 20 foliose, 6 fruticose,
and one squamulose morphologies (Fig. 3b). The foliose/crustose
thallus ratios were from 0.05 to 1 and decreased as follows: CHS
35 < SP 11 < HET 54a < PM 72 < BEX < EPC 74 < EPC 63 < EPC 08.
Biological richness and abundance (sum of frequencies) showed
a high heterogeneity among the study sites: from 13 to 35 species
encountered by individual site and the abundances ranged between
1.87 and 3.77 (Table 4). SP 11, PM 72, and CHS 35 showed a
Fig. 3. Lichen diversity found in the eight study sites: average abundance of each lichen species (a) and relative proportion of each type of morphology (b).
Summary of main ecological, bioindication indices, and values of the six environmental variable from Wirth, 2010 of each plotting area.
study site ecological indices bioindication
Wirth, 2010′s environmental indices (%)
IAP LDV light temperature continentality humidity pH eutrophication
SP 11 35 2.85 4.43 241 57 55.8 52.3 44.8 26.3 32.6 43.2
HET 54a 33 3.77 4.30 263 75 62.7 51.1 42.9 21.8 37.6 47.5
EPC 74 30 3.60 4.27 227 72 76.4 50.1 45.8 18.5 36.5 46.0
PM 72 26 2.02 3.71 159 40 69.0 50.2 38.5 26.2 25.8 38.6
EPC 63 25 3.31 3.66 157 66 75.1 50.0 41.9 26.6 37.1 51.8
CHS 35 23 1.87 3.52 137 37 45.2 51.3 39.1 37.2 25.9 23.0
BEX 20 3.20 3.02 117 64 64.5 49.9 50.0 14.1 37.0 57.1
EPC 08 13 2.17 3.16 94 43 77.3 50.0 50.0 15.6 32.4 56.5
relatively low lichens abundance for a same range of richness (rich-
ness/abundance ratio from 12.3 to 12.9) compared to the other sites
(ratio from 6.0 to 8.8). The Shannon index, ranged between 3.02 and
4.43. It followed the lichen diversity values with the exception of
BEX site, which may be due to a higher abundance (Table 4).
The main lichen communities observed in the study sites were
commonly found in France (Coste, 2001; van van Haluwyn and
Lerond, 1993; van Haluwyn et al., 2009): Leprarion incanae Almborn
1948 (except in BEX and PM 72), including sciaphilous species (Lep-
raria incana), and Lecanorion carpinae (Ochsn.) Barkm, 1958 (except
in EPC 63 and CHS 35), including heliophilous, nitrophilous and
toxitolerant species (such as Lecanora carpinea,Lecanora chlarotera,
and Lecidella elaeochroma). Parmelion acetabuli Barkman 1958 was
found in four sites (BEX, HET 54a, EPC 08, and PM 72), including
Parmelia sulcata,Melanelixia glabratula, as well as Melanohalea exas-
peratula and Physcia adscendens, mainly heliophilous and slightly
neutrophilous and toxitolerant species. Other nitrophobous and
poleophobous communities were found locally: Graphidion scriptae
Oschner 1928 (with Arthonia,Graphis,Enterographa and Opegrapha;
in HET 54a and CHS 35), Cladonion coniocraeae Duvigneaud ex
James, Hawksworth & Rose, 1977 (with Cladonia ﬁmbriata; in EPC
08 and PM 72), and Calicion viridis ˇ
Cernh. & Hadaˇ
c 1944 (with
Chrysothrix candelaris; in BEX).
The sampling procedure, including both hardwood and conifer
trees as far as possible (Table 1), attempted to reduce the tree bark
inﬂuence by limiting to only sample the main representative tree
species in each site (i.e., ﬁr in SP11, spruce in EPC 63, beech for HET
54a, oak in CHS 35, etc.), and thus, the lichen communities adapted
to these tree species. We collected lichen samples on a total of 21
different tree species, from 3 to 9 by site. Considering dominant
tree species (n ≥5, Fig. 4), lichen richness observed on hardwood
trees was usually greater compared to richness on conifers, except
for Abies: p < 0.05 (Student test). Fraxinus was the tree species
with the greater lichen richness (9.6 species on average). Also, the
lichen communities found on deciduous trees (Lecanorion carpinae
and Parmelion acetabuli associated with other foliose and fruti-
cose species) differed from those on conifers (generally Leprarion
3.1.2. Bioindication indices
The highest IAP (>200) were found in HET 54a, SP 11, and EPC 74,
while EPC 08 and BEX showed the lowest values (<120), following
asimilar trend as lichen richness (Table 4). The different sampling
and/or calculation methods may limit the data comparison (Scerbo
et al., 1999). Lichen diversity values were also highest (>70) in HET
54a and EPC 74, but the lowest values (≤40) were for the two West-
ern stations (CHS 35 and PM 72), following the lichen abundance
3.1.3. Ecological features
For each lichen species, we studied ecological features through
six environmental parameters described by Wirth (2010): light,
temperature, continentality, humidity, pH, and eutrophication.
When ecological data were absent (i.e., for 25 species), we
used data from Nimis and Martellos (2008) database, as well
as other references (Clauzade and Roux, 1985; Dobson, 2011;
Fig. 4. Lichen diversity by tree-support species (n indicates the number of individuals for each tree genus).
Smith et al., 2009; van Haluwyn and Lerond, 1993). An aver-
age ecological value of each parameter was calculated for each
station based on individual value and frequency of each lichen
species. To better homogenize the indices between these differ-
ent references and to reduce the wide ranges (generally nine
levels are reported by Wirth, 2010), we introduced a new scale
of three levels (e.g., xerophytic/mesophytic/hygrophytic species,
acid/neutral/basic substrate pH, etc.). The results were expressed
using the frequency of each lichen species (Table 4).
The most important gradient were found for eutrophication
(from low, i.e., CHS 35 and PM 72, to moderate eutrophic species,
i.e., BEX and EPC 08) and light (with high proportions of helio-
philous species, i.e., EPC 08, EPC 74, and EPC 63, and species
with moderate light afﬁnity, i.e., CHS 35 and SP 11). In contrast,
mesophytic species were dominant indicating a low difference in
temperature among sites. On overall, lichen species were, on aver-
age, mostly acidophilic, xerophilic and moderately oceanic in all
3.2. Coupling ecological and biogeochemical approaches
To determine the resistance or sensitivity of each lichen species
to metal atmospheric pollution, we performed multivariate sta-
tistical analyses including the three diversity variables previously
studied (lichen richness, lichen abundance, and Shannon index), the
six ecological parameters mentioned above, the two bioindication
indices (IAP and LDV), and metal bioaccumulation data measured
in foliose lichen species (i.e., Xanthoria parietina,Parmelia sulcata,
or Hypogymnia physodes) estimated using the sum of enrichment
factors (EF) for 17 metals (Al, As, Cd, Co, Cr, Cs, Cu, Fe, Mn, Ni, Pb,
Sb, Sn, Sr, Ti, V, and Zn; see Agnan et al. (2015)). A PCA was then
performed and the ﬁrst two components (81% of the data variance)
were represented (Fig. 5).
The ﬁrst component (45% of the data variance) was inﬂuenced by
lichen abundance and LDV with negative scores. It was associated to
lichen species living on basic bark and eutrophic, continental, and
bright environments, as illustrated by EPC 74, BEX, HET 54a, and
EPC 63 sites. The positive scores were characterized by hydrophilic
species and metal EF data from bioaccumulation in lichen, inﬂu-
encing the two Western sites (CHS 35 and PM 72). The second
component (36% of the data variance) grouped the two diversity
indices (lichen richness and Shannon index), as well as IAP and
lichen species living in warmer environments. The temperature
could not explain this component due to the lack of ecological con-
trast in the study sites. This component distinguished SP 11 and
HET 54a with positive scores, and EPC 08, and to a lesser extent
BEX, with negative scores.
A CCA was performed on metal bioaccumulation data and
lichen species frequencies found for each study site (Fig. 6a,b).
This method was already used for lichen sensitivity to nitrogen by
Glavich and Geiser (2008). Only lichen species presented in at least
two different study sites were included in the CCA. We added in
the analysis the sum of EF of the 17 metals previously cited (Agnan
et al., 2014) and the two bioindication indices (IAP and LDV). The
IAP was explained by the ﬁrst axis (26% of the data variance), while
the second axis (21% of the data variance) evidenced an opposite
pattern between LDV and EF (Fig. 6a). Each lichen species was repre-
sented by a three letter code on Fig. 6b (see Table 5 for the species
correspondence). Since IAP was a diversity index (Fig. 5), it was
proved difﬁcult to classify the lichen species following the ﬁrst axis.
Using the EF position in the ﬁrst plot as factor of metal pollution
(Fig. 6a), however, we determined the degree of metal inﬂuence for
each lichen species depending on the position of the species in the
second plot (Fig. 6b). To scale this inﬂuence, we applied a geomet-
ric rotation using EF as the new y axis (y’). The rotated coordinates
allowed differentiation of sensitive vs resistant species based on the
EF values (i.e., projection on the EF gradient, y= 2.28 x;Fig. 6b). The
lowest and negative y’ indicated a resistant species to metal atmo-
spheric pollution and the highest and positive y’ a sensitive species
(Table 5). Given the range of y’ values of 3 (between −1.5 to +1.5),
we determined three groups of identical ranges as follows: y’ < −0.5
for resistant species, −0.5 < y’ < 0.5 for intermediate species, y’ > 0.5
for sensitive species. The list of resistant species included various
crustose lichens, while only two crustose species were present in
the sensitive list (Pertusaria coccodes and Caloplaca ferruginea). Two
foliose (Melanohalea exasperatula and Physcia tenella) and one fruti-
cose (Cladonia ﬁmbriata) species, however, were found as resistant
species. The number of sites where each lichen species was present
was given as conﬁdence information of y’.
4.1. Lichen diversity and communities
The diversity of corticose lichen species observed in the eight
forest study sites was generally lower on coniferous trees com-
pared to hardwood trees (Fig. 4), conﬁrming literature observations
(Selva, 1994). Lichen communities were likewise different between
Fig. 5. Principal component analysis (PCA) including ecological characteristics (normal), ecological indices (italic), bioindication indices (bold), and the sum of enrichment
factors of 17 metals (EF, bold and italic).
Fig. 6. Canonical correspondence analysis based on frequency of the 43 main lichen species (presented in more than two different sites, with a three letter code, see Table 4
for the species correspondence), bioindication (IAP and LDV) and bioaccumulation (sum of enrichment factors [EF] of 17 metals) indices for each study site.
these two types of trees with mostly sciaphilous communities on
conifers and heliophilous species on hardwood trees (e.g., Leprarion
incanae vs Lecanorion carpinae in EPC 74, respectively). Our sam-
pling method in open areas bordering forests allowed therefore
maximizing lichen diversity and communities: both sciaphilous
and heliophilous lichens were found as dominant species (Lepraria
incana and Parmelia sulcata, respectively; Fig. 3).
Overall, the lichen diversity observed in the study sites was
high. The number of lichen species was in the same range as those
observed in other European forests: e.g., in Italy (Giordani, 2007),
cnik et al., 2008), or Portugal (Pinho et al., 2004);
the Shannon index, however, showed higher values compared to
other European and North American forested sites (Mulligan, 2009;
Peterson and McCune, 2001). But, this range was higher than in
boreal environments (Kuusinen, 1996), probably in relation to spe-
ciﬁc climate conditions in cold regions. Indeed, the diversity data
from the literature are not always comparable since the sampling
methods used can sometimes lead to discrepancies between the
observations (e.g., Kuusinen and Siitonen, 1998; Selva, 1994).
Based on the indices of Nimis and Martellos (2008), 12% of the
overall taxa were pioneer species, with the maximum proportion
for BEX and SP 11 (25 and 20%, respectively) and the minimum
for CHS 35 (4%). In BEX, two common lichen species (Lecanora
chlarotera and Lecidella elaeochroma) were responsible for 98% of
the pioneer frequency, but these species can also be found in
non-pioneer environments (Pirintsos et al., 1995). The pioneer fre-
quency was not directly positively correlated with lichen richness
(Table 4), as suggested by Selva (1994). This can be explained either
by our sampling protocol in open ﬁeld limiting forest, or by inad-
equate Nimis and Martellos (2008) pioneer index applied in our
Differences were, conversely, observed among study sites
regarding ecological characteristics. For example, SP 11 and EPC 08
showed both nitrophilous and poleotolerant communities, while
nitrophobous species were found in CHS 35 and HET 54a. This
agreed with observations in atmospheric deposition sometimes
different from modeled estimates, particularly under-estimated in
the Pyrenees (SP 11) and over-estimated in the Armorican Mas-
List of resistant, intermediate, and sensitive lichen species relative to atmospheric metal pollution based on a bioaccumulation–lichen diversity coupling method. y’ value
indicates the new scale of lichen resistance/sensitivity to metals. The number of sites where each lichen species was present gives a conﬁdence information of y’.
lichen species number of sites code y’
resistant species Lecanactis subabietina 3 Lsu −1.442
Pertusaria leioplaca 2 Pli −1.402
Pertusaria albescens 4 Pal −1.093
Graphis scripta 3 Gsc −0.919
Cladonia ﬁmbriata 4 Cﬁ −0.893
Melanohalea exasperatula 2 Meu −0.854
Dendrographa decolorans 3 Dde −0.799
Ochrolechia pallescens subsp. parella 2 Opp −0.781
Ochrolechia androgyna 2 Oan −0.656
Pertusaria amara 3 Paa −0.620
Lepraria incana 8 Lic −0.615
Lecanora allophana 2 Lal −0.608
Physcia tenella 2Pte −0.555
Calicium salicinum 2 Csa −0.551
Acrocordia gemmata 4 Age −0.537
Schismatomma cretaceum 2 Scr −0.525
Arthonia radiata 4 Ara −0.513
intermediate species Lecidella elaeochroma 4 Lel −0.494
Chrysothrix candelaris 7 Cca −0.482
Lecanora chlarotera 7 Lch −0.428
Melanelixia glabratula 6 Mgl −0.389
Lecanora conizaeoides 3 Lcn −0.284
Lecanora expallens 2 Lex −0.234
Parmelia sulcata 8 Psl −0.187
Lecanora argentata 4 Lar −0.021
Ochrolechia turneri 2 Otu −0.018
Amandinea punctata 5 Apu 0.066
Lecanora barkmaniana 2 Lba 0.080
Buellia disciformis 2 Bdi 0.124
Lecanora carpinea 3 Lca 0.170
Parmelina carporrhizans 2 Pca 0.204
Xanthoria parietina 5 Xpa 0.244
Phlyctis argena 4 Par 0.493
sensitive species Pleurosticta acetabulum 2 Pac 0.519
Caloplaca ferruginea 2 Cfe 0.524
Pseudevernia furfuracea 2 Pfu 0.590
Hypogymnia physodes 5 Hph 0.706
Evernia prunastri 6 Epr 0.732
Usnea sp. 2 Usn 0.739
Physcia adscendens 4 Pad 0.771
Ramalina farinacea 4 Rfr 0.824
Pertusaria coccodes 3 Pco 1.256
Physconia distorta 2 Pdi 1.405
sif (CHS 35; Boutin et al., 2015;Pascaud et al., 2016). Even though
no obvious correlation was observed with lichen richness, lichen
abundance, or foliose/crustose thallus ratio, lichen communities
agreed with the ecological features described by Wirth (2010):
e.g., CHS 35 had a low percentage of eutrophic species, unlike EPC
08. These ecological observations were therefore a complementary
description to assess environmental quality that cannot be illus-
trated by lichen richness or abundance only.
Results of bioindication indices showed that IAP were largely
higher than data from French urban areas (Gombert et al., 2004),
and LDV were generally in the upper range compared to other forest
sites in Europe (Giordani, 2007; Pinho et al., 2004; Poliˇ
cnik et al.,
2008). These indices were closely related to lichen richness and
lichen abundance, respectively (Table 4), which was supported by
the PCA results (Fig. 5). This is most likely due to their calcula-
tion method: only frequencies were used in LDV whereas Qi (i.e.,
the number of companion species, largely inﬂuenced by lichen
diversity) is considered in IAP. Thereby, the difference of results
between IAP and LDV, already observed by Poliˇ
cnik et al. (2008),
can be attributed to the difference between lichen richness and
lichen abundance strongly highlighted with the Northwestern sites
(PM 72 and CHS 35) and SP 11, showing a high number of lichen
species weakly abundant. Each index was mainly inﬂuenced by one
principal component (Fig. 5): axis 1 for LDV (45% of the data vari-
ance) and axis 2 for IAP (36% of the data variance). Based on lichen
ecological features (Nimis and Martellos, 2008), the signs of envi-
ronmental alteration (e.g., acid or poor nutrient environment) were
mainly inﬂuenced by the positive scores of the ﬁrst component,
i.e., opposed to the LDV. It is likely that IAP, and thus lichen diver-
sity, were mostly driven by climate variable (temperature) despite a
low gradient of temperature among lichen species. The sites PM 72
and CHS 35, both positively inﬂuenced by the ﬁrst axis, may either
reﬂect an environmental alteration (i.e., more acid conditions), or
be driven by the continentality–humidity axis due to their location
with Atlantic inﬂuence.
4.2. Resistance and sensitivity of lichen species to metal
As observed in the PCA (Fig. 5), the LDV was opposed to the sum
of metal enrichment factors in the axis 1 vs axis 2 plot. This implies
that, in addition to the response toward the general alteration of
environment, this index better responds to metal pollution as well.
Indeed, the three lowest LDV were observed in CHS 35, PM 72, and
EPC 08 (positive scores of the ﬁrst axis of the PCA and negative
scores of the second axis), that correspond to the highest EF and
metal deposition (as observed in EPC 08; Gandois et al., 2010b).
The northeastern France is impacted by various activities (local
industries, metallurgy, and mining), while both energy and metal-
lurgy may explain such contamination in the northwestern France
(already observed in upper horizons; Hernandez et al., 2003). This
may be the dominant inﬂuence for CHS 35 and PM 72 in the PCA
toward other environmental variables. Thus, it can be supposed
that metal pollution affects more lichen abundance (illustrated by
LDV) than lichen richness (IAP). This is in agreement with results
from Jeran et al. (2002), who had previously observed that IAP was
not a good index for metal pollution.
Based on the CCA, we evaluated the resistance or sensitiv-
ity of each lichen species to metal pollution (Fig. 6 and Table 5).
Very few literature observations, however, allowed supporting our
results: Cladonia ﬁmbriata (present in 4 sites, y’ = −0.893) is a well-
known species able to grow on cadmium, lead, and zinc enriched
substrates (Cuny et al., 2004; Tyler, 1989), whereas conversely,
Hypogymnia physodes (present in 5 sites, y’ = 0.706), is known as
a metal sensitive species, particularly for copper (Hauck and Zöller,
2003). To validate our results, we veriﬁed any correlations with
other pollutants: in both resistant and sensitive groups. There were
both acidophilic (e.g., Graphis scripta,Pertusaria albescens,Pertusaria
coccodes) and nitrophilic (Dendrographa decolorans,Physcia adscen-
dens,Physconia distorta;Gombert et al., 2004) species, as well as
both tolerant (Melanohalea exasperatula,Physcia adscendens) and
sensitive (Ochrolechia pallescens,Lecanora allophana,Physconia dis-
torta;Wirth, 1991) to SO2/NO2pollution species. This implies that
we cannot attribute the y’ values to sulfur and nitrogen pollution
inﬂuence, these elements being well known as major atmospheric
pollutants. In this way, our method allowed correct evaluation of
the inﬂuence of metal without other major disturbance. However,
organic pollutants also accumulated by lichens (Bajpai et al., 2010;
Harmens et al., 2013), were not investigated here. By applying the
frequencies of studied species to these indices, and comparing to
the enrichment factors from Agnan et al. (2015), we observed that
the four more polluted sites (i.e., HET 54a, EPC 08, CHS 35, and PM
72) as evidenced by bioindication, obtained negative scores (i.e.,
dominated by resistant lichen species), while several less contam-
inated sites (e.g., EPC 63 and EPC 74) obtained positive values (i.e.,
dominated by sensitive lichen species).
These preliminary data need to be completed and compared
with additional data from other European forest sites. Thus, it will
be possible to determine the maximum exposure of metal pollu-
tion without signiﬁcant harmful effects (also called critical load) as
already done for nitrogen (Geiser et al., 2010).
This study aimed to evaluate the resistance or sensitivity of
lichen species to atmospheric metal pollution. We performed
eight lichen plottings in French and Swiss forested sites, and
used different biomonitoring approaches (lichen richness, lichen
abundances, lichen community description, ecological features,
bioindication indices, as well as metal bioaccumulation) for a com-
plete environmental description. Each method provided its own
contribution to this investigation; similar results were demon-
strated by lichen communities and ecological features. Ninety-two
corticolous species were sampled, including 70% of crustose
lichens. The abundance was higher on hardwood trees compared to
conifers. The lichen diversity value (LDV) showed a better response
to both ecological disturbances (largely inﬂuenced by light and
nutrient conditions, such as eutrophication and pH) and metal pol-
lution compared to the index of atmospheric purity (IAP).
Using a multivariate approach coupling frequencies of each
lichen species and metal bioaccumulation data, we performed an
innovative scale of resistance/sensitivity to metals for the 43 more
frequent lichen species, distinguishing sensitive, intermediate, and
resistant species to metal pollution. To validate these results, we
compared to the few data available in the literature, and checked
any correlation with sensitivity to acid and nitrogen pollution. This
approach constitutes a ﬁrst insight into the investigation of resis-
tance and sensitivity of lichen species to metals in open forested
sites far from local pollution sources, which should be enhanced by
results with data from other European forests in future researches.
This project beneﬁted from ﬁnancial support number
1062C0019 from ADEME (French Agency for Environment).
The authors thank Clother Coste for his help in lichen determina-
tion. Yannick Agnan was funded with ADEME fellowship. Thanks
to four anonymous reviewers for their relevant comments that
improved this manuscript.
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