Considerable work has been done on the effects of air pollutants on lichens and bryophytes. Since lichens are slow-growing and long-living organisms with a special ability to accumulate substances from their environment, they are susceptible to many pollutants present in the atmosphere or brought down in the rain.
This sensitivity is, heightened by the fact that, unlike tropophytes, they never shed their toxin laden parts. Similarly, bryophytes, especially mosses, with their delicate and uncuticularized plant body, seem to have a marked capacity for absorbing and accumulating pollugenic substances from the environment.
High concentrations of heavy metals are phytotoxic and bryophytes and lichens can effectively absorb and retain these ions. The profusely branched and ramifying pleurocarpous mosses and certain acrocarpous mosses entrap and absorb particulate matter more efficiently than the erect and unbranched mosses. Cladonia deformis, Lecanora muralis and Peltigera rufescens seem to have a special affinity for iron, wheres Parmelia tinctorium, Verrucaria vigrescens and Stereocaulon nanodes are highly tolerant to zinc.
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Dicranella varia and Hylocomium splendens grow in shale spoil heaps and copper smelter soils (habitats rich in lead) whereas Cladonia sp. and Pohlia nutans are the characteristic species in copper swamps. Many lichens and bryophytes act as sinks for heavy metals, accumulating the metal cations by many times the concentration found in their associated substrates; such plants may play an important role in trapping mobilized metals and removing them from surface ninoff in disturbed ecosystems (Rao et al., 1977). Indeed, these lower plants seem to possess physical characteristics which make them ideal monitors of air quality.
Since lichens and bryophytes (especially the epiphytic ones) are outstanding among plants in their response to air pollution, they are generally considered to be reliable indicators of pollution. Certain species of these lower plants have been employed as bioassy organisms for estimating the content of such noxious gases as SO2 in air.
In fact, these organisms provide the simplest and the most economical tool for assessing an air pollution problem in time and space. The presence or absence of various lichens, along with a study of their morphological features, functional changes and developmental aspects has been well-correlated with SO2 content in the air (see Ferry et al., 1973).
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Lichens have in general been found to be very sensitive to virtually every kind of pollutant and more especially to pollutants rich in noxious gases, fluorides, heavy metals, radionuclides, agricultural chemicals and biocides.
Recent studies in England (see Ferry et al., 1973) have brought to light the fact that many areas which previously used to harbour diverse species of lichens are now virtually ‘lichen deserts’ since most of the lichen species have been exterminated by the increasing pollution generated by industrial and other activity in the vicinity of big cities.
For instance, Epping Forest, which used to have some 120 recorded species of lichens a century ago, has now only about 30, the rest being no longer seen. Some laboratory studies on the effects of SO2 on lichen physiology have also indicated its adverse effects on photosynthesis, respiration and soredial viability.
On the basis of laboratory and field studies it is now possible to estimate SO2 content in air simply by noting the species of lichens found in the particular locality. Thus, in England the lichens Parmelia saxatilis and P. fuliginosa did not occur on sand-stone walls above an annual average SO2 concentration of 60 ng/m3 and 45 p.g/m3 (Edwards, 1972).
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The methods of studying air pollution effects on lichens and bryophytes have been mostly phytosociological and ecophysiological. By means of these methods it is possible to relate the presence or absence of species, their numbers, frequency, coverage, and external and internal symptoms of injury to the degree of pollution vis-a-vis the purity of the atmosphere (see Rao and LeBlanc, 1967).
To obtain the basic phytosociological data, epiphytes of a single tree species are investigated at ecologically homogeneous sites in the polluted area. The “Index of Atmospheric Purity (IAP)” for each site is then calculated by means of the following formula.
where n is the total number of epiphyte species, f is the frequency-coverage score of each species expressed by a number on a numerical scale, and Q is the resistance factor or ecological index of each species determined by adding the number of its companion species present at all the investigated sites followed by dividing this total by the number of sites.
High IAP values represent rich epiphytic vegetation in relatively unpolluted areas whereas low IAP values indicate poor and depauperate epiphytic vegetation in polluted environments (LeBlanc et al, 1974).
In ecophysiological studies under field conditions, lichens and bryophytes are transplanted along with their substrates from unpolluted sites to ecologically similar but polluted sites.
After a certain known exposure period (4-12 months), the transplants from polluted sites are compared with those from non-polluted sites with respect to injury and other changes induced under the influence of pollution.
LeBlanc and Rao (1973) in their transplant studies observed that lichens that were transplanted in SO2-polluted areas showed abnormal microscopic features such as reduction in thallus thickness, deposition of a whitish, hydrophobic, acetone-soluble substance on thallus surface, and plasmolysis as well as chlorophyll degradation in the algal symbionts. Another noticeable effect was the failure of transplants to develop reproductive structures, vegetative as well as sexual, in the polluted areas.
On the basis of their laboratory studies, Rao and LeBlanc have shown that the algal symbiont of a lichen thallus is most vulnerable to SO2 pollution. They have proposed that the following chemical reactions take place dining hlorophyll degradation of the lichen algae under the influence of SO2:
SO2+H2O O H2SO22–
H2SO32– O HSO3–+H+
Chlorophyll-a+2H+ -> Phaeophytin-a+Mg2+
The mechanism of chlorophyll damage by hydrogen fluoride seems to be different from that by SO2. Fluoride perhaps combines with the Mg2+ in the chlorophyll molecule rendering it useless for photosynthesis. It has also been noted that the carotenoid pigments are resistant to destruction by SO2 as well as HF pollution.
Coal contains some amounts of mercury. When coal is burnt, this mercury is vaporized but some of it may condense on fly ash particles. Mercury from coal is thus released into the environment both as vapour and metal (e.g., as mercuric oxide) adhering to fly ash particles. The latter ultimately settle down at points some distance away from the source of discharge.
Huckabee (1973) has collected samples of fly ash, mosses, and other vegetation at varying distances from coalfly ash sources and has analyzed these for mercury content. He found that mosses (Dicranum scoparium and Polytrichum commune) had significantly higher mercury concentrations than the other plants collected at the same location.
The mercury content of mosses was found to depend on the distance from the coal smokestacks and ranged between about 1.1. ppm in polluted areas to 0.1 ppm in areas remote from the polluted zones. Some experiments using radioactively labelled mercury demonstrated that mosses take up and retain mercury to a greater extent than grasses. According to Huckabee, no source of regional mercury pollution is more important than the burning of mercury containing coal.
One of the most important aspects of the mercury pollution problem is the mercury accumulation by fish, both in terms of ecological effects and human consumption. Fish acquire and retain methyl mercury to a greater extent than other organisms and inorganic sediments. Species at different trophic levels are known to accumulate mercury at different rates.
For instance, crayfish, salamanders, and fish growing in the same stream accumulate methyl mercury at different rates. Zooplankton accumulates methyl-mercury faster than do insects. In contrast to methyl-mercury, fish concentrate fewer mercuric ions from mercuric nitrate than other biota. Even the mechanism of uptake and retention of inorganic mercury compounds by fish differs from that of organo- mercury compounds.
In aquatic habitats, fish and other organisms acquire body burdens of mercury either directly from the water via the gills or via the food-web. There is some indication that food-web uptake of mercury may account for a major part of the mercury body burden in fish.