The use of Lichens as a Pollution Indicator
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Bioindicators, also known as biomarkers, are tools that used in ecology, physiology, environmental microbiology and other disciplines, to detect strain and other environmental conditions surrounding organisms. USEPA stated that biological indicator species are unique environmental indicators as they offer a signal of the biological condition in a watershed as they reveal the pollution status from time to time. The presence of this species can give an early warning of pollution or degradation in an ecosystem as well as help sustain critical resources. Bioindicators differ from biomonitoring. Bioindicators are actually groups or types of biological resources that can be used for identification and qualitative determination of human generated environmental factors Tonneijk and Posthumus (1987) cited in Conti and Cecchetti (2001) while biomonitors are organism mainly used for the quatitative determination of and can be classified as being sensitive or accumulative (Conti and Cecchetti, 2001).
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Bioindicators are divided into two types. First, accumulation indicators which are store pollutants without any evident changes in their metabolisms. Another one is response indicators which is react with cell changes or visible symptoms of damage when taking up even small amounts of hazardous substances. The application of bioindicators in evaluating air pollution is much cheaper compared with other methods as Fuga et al., (2008) noted that beside low cost, they also easy for sampling, and the possibility of monitoring wide areas. The application of plants as bioindicators has been conducted for many years to detect environmental changes (Mokhtar et al., 2006). According to Chandra and Sinha (2000) plants are highly effective and sensitive tools for determining and predicting environmental stresses. Mosses, lichens and fungi are the examples type of plants that commonly used to indicate the environmental changes. However a lichen species was more resilient than a moss species in two exposure experiments investigating trace metal uptake involving transplants into urban environments (Tretiach et al., 2007).
Lichens consists of a fungus, known as the mycobiont, and a photosynthetic organism, a green alga or cyanobacteria species, the photobiont. Both of them are symbiosis as they rely each other where the algae and fungi give benefits each other (Hawksworth and Rose, 1976). The symbiotic action between the mycobiont (-s) and the photobiont (-s), form thallus which the body of lichens through a process of symbiogenesis (Margulis and Barreno, 2003) are stable “micro-ecosystems” as a result symbiosis process make these non-linear properties to be particularly effective ‘early warning indicators’ of changes dues to the impact of humans on ecosystems (Barreno, 2003). Poikolainen (2004) stated that the fungal component usually an Ascomycetes fungus, and a green alga (Chlorophyceae) and/or blue-green alga (Cyanobacteriae) is the algal component.
Nearly 19% of all fungi are lichenized (Lutzoni et al., 2001; Hawksworth et al., 1995). More than 98% of lichenized fungal species belong to phylum Ascomycota, others are from orders of phylum Basidiomycota and some to Mitosporic fungi (Hawksworth et al., 1995; Tehler, 1996). Most of the photobiont partner is form by green alga (Chlorophyta; Lewis and McCourt, 2004), only about 10% with cyanobacteria, and 3% with both green alga and cyanobacteria (Hawksworth et al., 1995; Honegger, 1996).
The fungal component is responsible for taking up water and minerals, and the algal component, which grows in the middle of the fungal mycelia, for photosynthesis in order to feed both partners as it has chlorophyll that is able to consume sunlight, produce essential nutrients. Most lichen species obtain their nutrients from wet and dry deposition (Garty, 1993). Some mycobionts can also change their photosynthesizing partner from green alga to cyanobacterium and vice versa and this leads to changes in thallus morphology (Oksanen, 2006). This behavior was suggested to be due to an environmental adaptation and related to ecological compatibility of the photobiont (Honegger, 1996; Stenroos et al, 2003).
The classification of lichens depends on thallus structure. There are three growth forms of lichens which are Crustose (crusty), foliose (leafy) and fructicose (shrubby) (Swinscow and Krog, 1988). Growth form irregular the degree of physical contact and orientation of the lichen with its substrate as well as the amount of continuous surface area exposed to airborne deposition therefore it should have a direct impact on both the interception and uptake of airborne and substrate available elements by lichens (Samuel et al, 2002).
Basically lichens can be found in terrestrial habitats, although a few can survive constantly below the surface of water, such as Peltigera hydrothyria. They can live on bark (epiphytic), rocks (epilithic), or soil (terricolous) and can even grow within the upper portion of rocks (endolithic), particularly in exposed limestones and sandstones. In the tropics they can also colonize leaves (foliicolous) (Nash, 2008). Lichens often grow in habitats with extreme light, dryness, or temperature, which are less favorable or unsuitable for higher plants (Vrablikova et al., 2006). Although lichens are attached to the bark or penetrate a short distance, they are not parasite which is not entering the inner bark where food is transported, and hence do not consume the tree of nourishment. Lichens depend on mineral nutrients from wet and dry deposition on the plant surface in the form of soluble salts and particles, for their growth and metabolism (Loppi and Pirintsos, 2003).
Lichens can extend the extreme condition, slow-growing, organisms that maintain a fairly uniform morphology in time and are highly dependent on the surrounding for nutrients (Loppi and Pirintsos, 2003). Several researchers such as Chiarenzelli et al. (1997) have studied the accumulation of heavy metals in Arctic tundra ecosystem at the Otter Lake, Northwest Territories, Canada.
Lichens as bioindicators
Since lichens are the most widely used biomonitors in terrestrial environments (Nimis et al., 2002), therefore they can detect and monitor a lots of pollutants such as SO2, HF, metals, high nitrogen deposition, organic pollutants and radionuclides. Most studies using lichens as indicator species of air pollution particularly acid rain, fertilizers, sulphur and nitrogen oxides, and metals, has been documented in thousands of scientific papers (Henderson, 2000).
Hawksworth and Rose (1976) reported that in the early 1860’s lichens were recognized as potential indicators of air pollution in Britain and Europe. Since then, lichens have played prominent roles in air pollution studies throughout the world because of their sensitivity to different gaseous pollutants, particularly sulfur dioxide. More than a century ago William Nylander (1866), a European scientist, found that lichens in the countryside around Paris were not found inside the city. He investigated the situation and found out that the lichens had been rapidly diminishing, killed by pollutants. Since of those findings, an extensive research have been conducted in many areas (Barkman, 1958; De Wit, 1976; Hawksworth, 1971) cited in Wolterbeek et al., (2003).
Other than that, several researches have studied the relationship between lichens and trace elements in different geographic area (Loppi and Bonini, 2000; Garty, 2001; Carreras and Pignata, 2002; Yenisoy-Karakas and Tuncel, 2004; Conti and Cecchetti, 2001; Bergamaschi et al., 2004). Hundreds of studies have been published on the effects of sulphur dioxide, nitrogen compounds, ozone, heavy metals and other atmospheric pollutants on the morphology and physiology of lichens since the 1950’s (Richardson, 1992; Garty, 2000). These studies have primarily been experimental.
Lichens have also been employed among other things as accumulation indicators of heavy metals (Freitas, 1994) as well as sulphur and nitrogen compounds (Søchting, 1995) derived from industrial activities and power production in a countless number of studies carried out in the surroundings of emission sources. Therefore research related to the pollution based on lichen as an indicators being used widely. This is because there were certain characteristics on lichens that make them an excellent bioindicators for determining the presence of the primary pollutants such as sulphur dioxide (SO2), nitrogen dioxide (NO2), fluoride, acid precipitation, ozone and metals (Hutchinson et al., 1996). Several regional and even national surveys have been carried out on the relationship between the occurrence of epiphytic lichen and atmospheric pollutants (Poikolainen et al., 2000).
Other than that, Soderstrom (1988), Lesica et al., (1991), Esseen et al., (1997) cited in Humphrey et al., (2002) supported that epiphytic bryophytes and lichens are important components of biological diversity in natural boreal and temperate forests. Most of species of lichen have a wide geographical distribution, which allows for a study of pollution covering wide areas and its high capacity to accumulate metals Burton (1986) cited in Mokhtar et al., (2006). Hutchinson et al., (1996) stated that lichens do not have seasonal variations and therefore accumulation of pollutants can occur all years. They cover 8 % of the land surface, including some of the most extreme environments on Earth Larson (1987) cited in Backor and Loppi (2009). Lichens and mosses usually have considerable longevity, which led to their use as long-term integrators of atmospheric deposition (Sloof, 1993).
Lichens are very sensitive to disruption in naturality resulting from air pollutants, primarily sulfur dioxide and heavy metals (Nimis et al., 2002). Their unique characteristic which is do not have outer impermeable layer of tissue to prevent gases and particles that affect their metabolism. This characteristic also indicate that small particles in the atmosphere can be absorbed make them most important bioindicator (Bennet, 2006). Besides that, because of the weakness of its cuticles enables moist air to be absorbed through its surface. Metals which are absorbed along with the moist air are dissolved in it (Mokhtar et al., 2006). Trace metals from air borne particles can be absorbed by lichen and then accumulate and saturated the metals. This is because their structure and anatomy (Hutchinson et al, 1996) make them able to uptake the pollutants.
The diminishing of some lichen diversity due to the increasing in air pollution and environmental stress (Svoboda, 2010) indicate that the environmental condition in that region is polluted. Pollution forms such as forest fragmentation (Fritz et al., 2008; Hedenas and Ericson, 2008; Ranius et al., 2008) give premium utility as indicators of naturalist.
Lichens can be considered and analyzed in terms of their morphology, histology, ecology and physiology, in short or long-term periods of time (Ahmadjian and Hale, 1973). Several researches Garty et al., (1998) has been studied on the lichen’s biological performance such as measurements on growth rate, productivity, reproductive capacity, deformity, discoloration, chlorophyll content, membrane integrity, respiratory activity, ionic content, geographical occurrence, substrate-related distributional limitations, or water relations. Habitat degradation and loss (Groom et al., 2006), habitat fragmentation (Bergamini et al., 2005), overexploitation (Upreti et al., 2005), species invasions (La Greca and Stutzman, 2006), and climate change is the main threats to the biodiversity of lichens.
For example, overbrowsing of the Cladonia heath by increasing reindeer populations in Scandinavia and Alaska has long been recognised as an important factor causing the severe decline of lichens, which might become a serious problem in reindeer husbandry (Suominen and Olfosson, 2000). Climate change is likely to have dramatic effects on distribution and abundance of lichen populations (Ellis and Coppins, 2007; Ellis et al., 2007). Yet another threat which is specific to lichens and other poikilohydric cryptogams is air pollution, which has led to the severe decline of numerous species throughout Central Europe (Nimis et al., 2002).
The structure and characteristics of lichens play important role as they served as the early warning to the pollution. Otnyukova et al. (2007) identifies a relationship between deposition, abnormal morphology in Usnea and tissue chemistry, providing an early indication of forest decline.
However the biological scaling, the interpretation of “symptom mapping” is often rather difficult Seaward (1976) cited in Wolterbeek et al., (2003). Due to the increase in size may reflect an increase in reproductive capacity but may also only show an abnormality of shape or form of the central parts of the thallus. Another reason is the discoloration might be associated to the general deteriorate but also reflect injuries from other sources such as insect, diseases or chemical sprays.
Last but not least, changes in the distribution of species may reflect changes in air pollution (Barkman, 1958; De Wit, 1976; Henderson-Sellers and Seaward, 1979 cited in Wolterbeek et al.,2003), but may also be associated to other environmental parameters (Henderson-Sellers and Seaward, 1979), such as changes in humidity or changes in the availability of preferent substrates (De Wit, 1976; Manning and Feder, 1980).
Determining concentrations of heavy metals in the environment is an important part of understanding biochemical processes and degree of ecosystem health (Schilling and Lehman, 2002). According to USEPA in Private Drinking Water Wells, heavy metals can be define as metallic elements with high density, such as, mercury chromium cadmium, arsenic, and lead. Even at low levels these metals can damage living things. Once heavy metals get into the environment, whether in small or large quantities, they cannot be completely eliminated. They cannot to break down or biodegrade and tend to build up in plants, animals, and people causing health concerns. This statement supported by other researchers such as Lenntech (2004) cited in Duruibe et al., (2007) which stated that any heavy metal is toxic or poisonous even at low concentration. However, their density is not the most concerns characteristics to be considered but their chemical properties (Duruibe et al., 2007) as they can cause significant impact to the environment. (Dembitsky, 2003) reported that these toxic substances contribute to a variety of toxic effects on living organisms by food chain as they enter into environment.
Source of metal pollution
Heavy metals occur in two ways; naturally or geological and anthropogenic activities such as industrial effluents, fuel production, mining, smelting processes, military operations, utilization of agricultural chemicals, small-scale industries (including battery production, metal products, metal smelting and cable coating industries), brick kilns and coal combustion (Zhen-Guo et al., 2002). This statement supported by Dembitsky (2003) which stated that these pathways are sources of heavy metal contamination. Other than that Zhang et al. (2009) noted that in nature, heavy metals are widely distributed in such ways such as water, soil, air and various forms of organisms at low concentration.
According to fairfaxcounty (2005) nowadays roadways and automobiles are considered to be one of the largest sources of heavy metals. The most common heavy metals released from road travel are zinc, copper, and lead, and at least 90 of the total metals in road runoff. However lead concentrations, consistently have been decreasing since leaded gasoline was stopped. Other than that, insignificant amounts of other metals, such as nickel and cadmium, are also found in road runoff and exhaust.
Transition metals are required by plants (Akbulut et al.,2008; Sofuoglu, et al.,2008). Wolterbeek et al. (2003) stated that the elements can be classified into macronutrients or micronutrients and as essential and non-essential. Some metals; cobalt (Co), copper (Cu), chromium (Cr), and nickel (Ni) are actually necessary for human in extremely small amounts (Zevenhoven and Kilpinen, 2001). However some elements for example mercury (Hg), cadmium (Cd), lead (Pb), chromium (Cr) and arsenic (As), can cause pollution and disrupt the environment when their accumulations exceed certain levels even at low concentration (Kennish, 1992).
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According to Jadia and Fulekar (2009) classified that “essential” metals which have benefits and naturally found are (Ca, Co, Cu, Fe, K, Mg, Mn, Mo, Na, Ni, Se, V and Zn) but can be toxic when excessive while “non essential” metals which are (Al, As, Au, Cd, Cr, Hg, Pb, Pd, Pt, Sb, Te, Tl and U) that can be highly toxic and can cause serious health damage when excessive uptake. Information given by Bowen (1979) cited in Wolterbeek et al. (2003) and Markert (1996) can be use for further general and more detailed data on metal essentiality, occurrence in soils and plants, toxicity and uptake. Plants have the ability to accumulate essential metals in different concentration for growth and development (Jadia and Fulekar, 2009). However plants also tend to accumulate the non-essential metals which have no known biological function (Djingova and Kuleff, 2000). Wolterbeek et al. (2003) found that because of the plant’s metabolically controlled maintenance of required levels of essential elements, lichens may show rather high base-line concentrations for especially the essential elements under low atmospheric availability conditions.
Because of their characteristics which is cannot to be brake down, it can cause direct toxicity by damaging cell structure (due to oxidative stress caused by reactive oxygen species) and inhibit a number of cytoplasmic enzymes when the accumulation inside plant cells above threshold or optimal level (Assche and Clijsters, 1990). Furthermore, it can cause indirect toxic effects by replacing essential nutrients at cation exchange sites in plants (Taiz and Zeiger, 2002). Baker (1981) cited in Jadia and Fulekar (2009) suggested, that some plants have evolved to tolerate the presence of large amounts of metals in their environment by the following three ways; exclution, inclution and bioaccumulation.
Lichens and excess of heavy metals
Permeability of the plasma membrane of lichens may alter by metals, leading to leakage of ions like potassium and other solutes (Grunsveld and Clijsters, 1994). Biomonitors by lichens at specific area has weakness as their responses could be varying from those observed in the laboratory. This is due to the fact that the lichens are exposed not only to a single pollutant but to other mixtures of pollutants which are also affected by different meteorological conditions (Carreras et al., 2005). Although biomonitoring program on atmospheric heavy metals and progress in this field has been already reviewed throughout the world (Garty, 2001; Bargagli and Mikhailova, 2002) but in the last few years there has been researched focused on the physiological and biochemical effects of heavy metal accumulation in lichens (Backor and Loppi, 2009).
There are many studies has been documented regarding the effectiveness of lichens in intercepting particles not only from the atmosphere but also from substrate (Loppi et al., 1999; Pirintsos et al., 2006). These particles may be precipitated onto the lichen surface or trapped in the intercellular spaces of the medulla (Garty et al., 1979 cited in Backor and Loppi, 2009) and remain unchanged for a long period. Lichen can be damage caused by the presence of other gaseous or particulate pollutants in the environment which can interrupt the physiological processes involved in the accumulation of heavy metals (Carreras et al., 2005). However Backor and Loppi (2009) found that because of the particles deposited and remain unaltered, lichens can retain and accumulation of heavy metals in quantities that exceed their physiological requirements.
Several studies have been conducted on the accumulation of heavy metals in many different species (Sawidis et al., 1995; Monaci et al., 1997; Scerbo et al., 1999; Loppi et al., 2002). Some common species such as genera Acarospora, Aspicilia, Lecanora, Lecidea, Porpidia, Rhizocarpon or Tremolecia (Purvis and Halls, 1996; BaÄkor and Fahselt, 2004) associated with heavy metal-rich substrates can tolerate metals and occur in both polluted and unpolluted areas. Other species, however, are restricted and have a lack connection in distribution reflecting the availability of suitable sites (Backor and Loppi, 2009).
Accumulation of heavy metals
Lichens tend to accumulate metals from airborne particles or from dissolved and suspended material (Wolterbeek et al.,2003). In general, five mechanisms have been put forward with regard to the absorption of metals in lichens; ion exchange, electrolyte sorption, intracellular absorption, entrapment of particles that contain metals, extracellular and hydrolysis as indicated by uptake studies with intact lichens (Richardson, 1995).
Accumulation of heavy metals in lichens is well documented (Bargagli and Mikhailova, 2002). Lichen tissue analysis probably arose from physiological investigations into the effect of heavy metals on lichen metabolism. Experimental work over the period from 1970-1985 by a group of lichen physiological ecologists, mainly based at Laurentian University in Sudbury, determined the precise relationship between elemental uptake, storage and lichen metabolism (Richardson and Nieboer, 1983; Richardson and Puckett, 1973). The role of lichens as accumulator the heavy metals of various trace elements in atmospheric deposition, and tissue analyses have effectively characterized their spatial and temporal deposition patterns has been studied by several researchers such as (Garty, 2001; Walker et al., 2003).
According to Baker (1983) cited in Contti and Cecchetti (2001) there are many factors that influence the accumulation of metals of plants such as the availability of elements, the characteristics of plants such as type of reproduction. Therefore, the degree of tolerance to heavy metals is characteristic of each lichen species (Carreras and Pignata, 2007). Neiboer et al. (1976) indicated a large range in the elemental uptake of lichens that varied according to elemental characteristics of the substrate and environmental factors, notably a ten-fold increase in metals in relation to distance from smelters. Metals and sulphur dioxide behave differently and are expected to have differing fallout patterns; however Rossbach et al. (1999) demonstrated a linear correlation between the element concentration in lichen material and the reciprocal of the distance from the emission sources. However, an adequate consideration of topography, substrate and meteorological conditions must be considered to explain variation.
Different species have different ability to absorb considerable amounts of heavy metals (Mokhtar et al., 2006). For example Pawlik-SkowroÅ„ska et al. (2006) found that apothecia of Lecanora polytropa accumulated Cu up to 1.3 %(d.m.), approximately 50 % of which was in an exchangeable form. Other than that a few studies have shown that foliose species generally had higher element loads than fruticose species when collected from the same site (Glenn et al., 1995; Gough et al., 1988; Lawrey and Hale, 1981 cited in Clair et al., 2002).
Various analytical techniques have been attempted by many researchers to increase knowledge on the bonding process which is interaction between lichen and metal. One of them is electron paramagnetic resonance. Although Conti and Cecchetti (2001) reported that there was lacking knowledge in understanding the entire process that responsible in the accumulation, but there were new approach recently been introduced which is by using microcalorimetric technique with the aim of obtaining enthalpic measurement data (Antonelli et al., 1998). As a result the trend established a good correlation between the metal bond and enthalpy values in metal uptake. For example trend for Evernia Prunastri Pb>> Zn> Cdâ‰ˆ Cuâ‰ˆ Cr (Conti and Cecchetti, 2001).
The algal partner has been reported to react more sensitively e.g., to acidic deposition and heavy metals, and to show varying accumulation of metals depending on the acidity of precipitation (Tarhanen et al., 1999). Sporadic desiccation of lichens may also have an effect on the accumulation and absorption of elements (Puckett, 1988). After a dry period, rainfall may result in appreciable washing off of particles and the exchange of cations bound on negatively charged exchange sites on the cell walls and plasma membranes of the cells (Bargagli, 1998).
The rapid, exchangeable process of metal binding to cell walls in metal uptake by lichens has been extensively studied in the laboratory (Brown, 1976; Goyal and Seaward, 1982; Nieboer and Richardson, 1981; Nieboer et al., 1978 cited in Wolterbeek et al., 2003). Several studies have been conducted to measure the deposition of atmospheric in lichen. France and Coquery (1996), for example measured deposition of atmospheric lead and compared them to lichen thalline contents from the high Arctic using flameless atomic adsorption spectrophotometry. They found that the concentration of 2 ug g-1 dry weights to be the lowest level found in lichen and established a latitudinal gradient in lead, with a clear pattern in decreasing Pb concentrations in lichens with increasing latitude throughout Canada.
Simple to complex methods for determining the concentration of ions in lichen tissue have been developed for decades and it is widely accepted that tissue concentrations in most lichen species shows a precise relationship with deposition of particulate and ambient air concentrations of gaseous pollutants (Seaward, 1992). Entrapment of particles contains metals which known as airborne pollutants accumulate in lichens by both wet and dry deposition (Nash, 1996). Wet deposition involves any kind of precipitation event that washes airborne aerosols and particles out of the air, while dry deposition involves the settling out of airborne gases or particles due to the increasing influence of gravity with decreasing wind speed (Knops et al., 1991). Another source of elements such as soil particles deposited by windblown on lichen thalli may be accumulated by some lichens. Smaller deposited particles may become trapped in the lichen thalli of species with large intercellular spaces (Collins and Farrar, 1978 cited in Clair et al., 2002.
However, when the concentration of metals is high enough to become toxic, they themselves cause damage to the lichen thalli. As a result, several physiological mechanisms of response to air pollutants in lichens are altered, and thus change their original sensitivity or tolerance to gaseous compounds like SO2, NOx, and O3 (Carreras and Pignata, 2007). These compounds affect the condition of lichens and thus reduce the capacity of lichens to accumulate and absorb elements from the atmosphere. Heavy metals have also been found to affect e.g., the permeability of the cell membranes of lichens (Tarhanen et al., 1996).
Heavy metal content in lichen thallus tends to alternate over time in phases of accumulation and subsequent release. Szczepaniak and Biziuk (2003) listed the factors that influence the metal absorption in lichens which are acid precipitation, geographical variations such as altitude, temporal changes for example seasonal variations, soil dust, local pollution sources, long-range transport.
Climatic factors probably play important role in the bioaccumulation of heavy metals, even if this yet unclear. Aptroot and van Herk (2007) provide increasing evidence that climate change is an important factor, partly based on evidence from the algal partner, much neglected in almost all studies, even though it is usually most sensitive to pollutants. The direction in which the pollutants are transported by wind direction is most surely fundamental in determining their main fallout point (Conti and Cecchetti, 2001). Accumulation of heavy metals in lichens is a dynamic process. Investigation on the effects of excess metals showed that lichens soaked into metal solutions accumulated metals quickly within a few hours. Observation on some metals such as copper showed maximum accumulation after 3- 6 h (Monnet et al., 2006).
In transplantation studies indicated that most lichens respond to changes in atmospheric heavy metals within a few months. It took 2-5 years to evaluate the elements in lichen thalli (Walther et al., 1990). Although it expected that the heavy metal content of lichens would increase as the time increase, but the situation is really much more complicated. Studies from Backor and Loppi (2009) showed that contents of several elements in transplanted lichens go up and down during the study period. This is because the contents of these elements are, at least partly, controlled by physiological processes and turnover mechanisms (Bergamaschi et al., 2007). Moreover, metals can be removed by rainwater which remove contaminating particles on the thallus surface (Brown and Brown, 1991) resulting in lower content during periods of rain and higher content in the dry season.
Other than that, the water-leachable fraction (deposited plus intercellular) is generally assumed to represent mainly metals originated from dry deposition, and the elements present in leachates show distinct temporal patterns with concentrations being usually higher in summer than in winter (Boonpragob and Nash, 1990). However, because lichens are more active metabolically when wet, winter months are suitable to growth and mineral uptake (Nash and Gries, 1995). Not only the water-leachable fraction contributes the uneven of the concentration, rainfall also richly contributes to the total element content of lichens (Knops et al., 1991). Both of the process could produce temporal differences in metal contents in spite of constant pollution loads, however the concentrations in lichen thalli mainly varies according to amounts of incident pollutants. These may produce a high amount of biological stress, thus alter element uptake (Bergamaschi et al., 2007).
Besides that, the time exposure of lichens transplanted into polluted areas also takes a count in the accumulation of elements process. This is because it influences the vitality of thalli and consequently the active processes of element uptake. Garty (2001) noted that in order to determine the minimal exposure required to produce significant change, the critical period of exposure remains unknown especially in the absence of time-studies. Short exposure times of 1-3 months are generally sufficient to affect transplanted lichens. When the exposure is longer, they become saturated with the elements, lose biomass, surface structures change and alter physiological performance (Bargagli and Mikhailova, 2002).
Other factors are the structures of lichen itself influence the metal uptake. Their physical characteristics such a surface structure, adhesiveness and water-holding capacity of thalli also affect metal accumulation in lichens (Brown and Beckett, 1985). Richardson (1995) observed that pores and holes on the cortex, which could trap particles or allow them to enter inside the thallus, in some species may be sealed by polymers, probably including lichenins and isolichenins. In polluted areas, the concentration of trace element in the peripheral (younger) of foliose lichens may be lower than central (older) part of the thallus because of the greater exposure times experienced (Bargagli et al., 1987 cited in Backor and Loppi, 2009). This comparison has been studied by (Loppi et al., 1997), where they used Flavoparmelia caperata thalli from an unpolluted area of central Italy.
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