1. LITERATURE REVIEW
1.3 INTRODUCTION TO LICHENS
1.3.4 Using lichens to monitor pollution
Lichens are particularly sensitive to anthropogenic activities which pollute the environment making them ideal biomarkers in assessing the effect of air pollution (chemical deposits), ozone depletion, metal contamination and nutrient concentration in the environment (FERRY et al., 1976; HAWKSWORTH and ROSE, 1976; ROSE and HAWKSWORTH, 1981; VOKOU et al., 1999). Lichens have no vascular system for conducting water or nutrients; thus they depend on atmospheric sources (e.g. fog and dew). However, major water resources for lichens often have much higher pollutant concentrations than precipitation. The lichens nutrient concentration mechanism also will concentrate pollutants (NASH, 2008). Unlike many vascular plants, lichens have no deciduous parts, and hence cannot avoid pollutant exposure by shedding such parts.
Furthermore, the lack of stomata and cuticle in lichens means aerosols may be absorbed over the entire thallus surface, thus lichens have little biological control over gas exchange, and air pollutant gases are assumed to readily diffuse down the photobiont layer (NASH, 2008). Although dehydration allows lichens to survive dry periods, it also concentrates solutions to the point that toxic concentrations may occur. Finally alteration of the symbiotic balance between the photobiont and mycobiont may readily lead to a breakdown of the lichen association (FEIGE and JENSEN, 1992; NASH, 2008).
14 Most lichen species live for decades or hundreds of years and a few even longer: thus as perennials they are subject to the cumulative effects of pollutants. Lichens can provide a long term reflection of local environmental conditions such as elemental composition of plant matrices and the fluxes of elements between biota, lithosphere and atmosphere. In general the elemental composition of plants reflects the chemical composition of their growth media such as soil, air and nutrient solutions (FARAGO, 1994). This was demonstrated in the study by DILLMAN (1992) where he reported a significant accumulation of elements such as Cu, Ca, Mg and K in lichens as a function of distance and direction from the refineries nearby.
Classic studies on lichens show that (sulphur dioxide) SO2 pollution resulted in the decline of many lichens especially around big cities such as London. Improvement of air quality and reinforcement of the clean air legislation in many European countries (e.g. 1956, 1968; clean air acts in UK) has resulted in dramatic reduction in pollution especially SO2
levels. This improvement was evident by the recovery of the lichen flora in the 1980‟s (VAN DOBBEN and DE BAKKER, 1996).
In recent years, agriculture-based pollution which includes ammonium depositions, herbicides and pesticides have been shown to have had a negative effect on lichen species richness (BROWN, 1992; ALSTRUP, 1992; MODENESI, 1993). The major sources of deposited atmospheric N are ammonia (NH3) and nitrogen oxides (NO and NO2) on a global scale, especially in Europe. Emission of reduced N, mainly resulting from livestock management and fertilizer application exceeds those of oxidised N (PITCAIRN et al., 2003). Acid and oxidising forms of N are broadly toxic and tend to reduce species richness (DAVIES et al., 2007). The supply of reactive nitrogen to global terrestrial systems has doubled and when released on land, they result in eutrophication of both fresh water and ground water, while emission to the atmosphere results in regional eutrophication and acidification, crop damage and impact on human health (GALLOWAY, 1998; ERISMAN et al., 2003).
15 Lichens are taxonomically diverse (TEHLER, 1996). Among lichens, oligotrophs (acidophytes) are adapted to environments with low nutrient availability compared to mesotrophs (neutrophytes) which require moderate N and eutrophs that thrive in nutrient rich environments (GEISER et al., 2010). Their relative dominance shifts with nutrient N deposition, allowing characterization of community effects and ecological harm (VAN HERK et al., 2003; MITCHELL et al., 2005; SPARRIUS, 2007; SUTTON et al., 2008). Due to their diversity, other species can survive increased N loads either by avoiding excessive assimilation (HYVÄRINEN and CRITTENDEN, 1998;
DAHLMAN et al., 2002) or an ability to store surplus N in the form of amino acids such as arginine (SILBERSTEIN et al., 1986). Such increases in tissue N concentrations are likely to cause an increased C demand, both to provide C skeletons for amino acid synthesis (NORDIN and NÄSHOLM, 1997) and energy for increased respiration (CHAPIN et al., 1987).
Several lichens have declined or even disappeared from habitats with elevated N levels (DAHLMAN et al., 2002) which might reflect uncontrolled N assimilation leading to toxic levels of NH4+ in the tissue (GAIO-OLIVEIRA et al., 2001). In the Netherlands, during the last 10 years an increase in nitrophytic lichen species, paralleled by a decrease in acidophytic ones, has occurred in areas with high cattle density (VAN DOBBEN and TER BRAAK, 1998). This phenomenon was especially apparent on acid-barker trees, on which nitrophytes were previously scarce or absent (VAN DOBBEN and TER BRAAK, 1998). A similar shift in species composition was also observed in the UK (WOSELEY and JAMES, 2002) and Switzerland (RUOSS and VONARBURG, 1995).
Comprehensive lichen mapping programmes for the British Isles has shown that some species have extended their ecological and geographical range by exploitation of acidified substrata (SEAWARD, 2004). Similar mapping programmes have been established in many other European countries (VOKOU et al., 1999).
British heaths consist of a large proportion of the remaining European heathlands, where mat-forming lichens are major vegetation components (FARRELL, 1989; HYVÄRINEN and CRITTENDEN, 1998). Genera such as Cladonia, Cetraria, Flavocetraria and
16 Sterecaulon grow as extensive carpets in the vegetation of oligotrophic tundra, sub-artic taiga and heath and ombrotrophic peabogs (RODWELL, 1991). They are important contributors to the functional ecology of these habitats, in terms of biomass and carbon storage (LANGE et al., 1998).
The need to study acid load in these fragile heathland ecosystems has prompted several studies. CRITTENDEN (1989) demonstrated the efficient uptake by Cladonia stellaris and Stereocaulon paschale of NH4+
and NO3-
delivered in rainfall (80%) and suggested that carpets of mat-forming lichens are significantly ombrotrophic in nature thus possess enormous potential as indicators of N deposition, particularly across large areas of remote northern terrain in which they are often abundant and for which measurement of N load are generally sparse. Mat- forming lichens occur typically in open situations where they intercept rainfall that is largely unmodified by overlapping plant canopies (HYVÄRINEN and CRITTENDEN, 1998). Recent studies by HOGAN et al. (2010a) revealed that mat- forming terricolous lichens provide coherent biomarkers for N enrichment. Interestingly, nitrogen enrichment also induces P-limitations in C. portentosa with attendant changes in chemical and physiological characteristics that could be used as a sensitive biomarker with which to detect low levels of N pollution (HOGAN et al., 2010a; 2010b).