## Air Quality and Lichens
**- A Literature Review** 

By Jenifer Hutchinson, Debbie May nard, and Linda Geiser USDA Forest Service, Pacific Northwest Region Air Resource Management Program Updated 16

- Dec

- 96



## Table of Contents

 Introduction

 Major Pollutants
o Sulfur dioxide
o Fluoride
o Acid Rain
o Photoch emical Toxin
o Metals
o
 Mechanisms of Accumulation
o Particulate Trapping
o Ion Exchange
o Accumulation of metals

 Phy siological Responses to Pollution
o Lichens as pollution monitors
o Effects of pollution on Lichens
o Sulphur dioxide
o Effects of acid rain

 Bark pH interactions

 Soil chemistry

interactions

 Other substrate interactions

 Acid fog

 Sensitivity

of lichens to sulfur dioxide vs. acid rain

 Nitrogen enriched acid rain

 Lichen structure and ph ysiology
o Effects of fluorine
o Effects of photochemical toxins
o Effects of metals

 Methods of Study
o Methods using naturally occurring lichens

 Distribution mapping

 Gradient studies
o Laboratory

measurements

 Elemental Anal ys is

 Measuring metals accumulated by lichens Phy siological responses

 Transplants

 Photograph y
 Air quality

monitoring studies using lichens in the Pacific Northwest
o West of the Cascade crest
o East of the Cascade crest

 Annotated bibliograph y



## Introduction Lichens are an integral and important component of Pacific Northwest ecosy stems. They contribute to nutrient and hydrological cycles (Pike 1978, Harper and Marble 1988, Ry chert and Skujins, 1974), biological diversity (McCune and Geiser 1997), and they are critical sources of forage, shelter and nesting material for man y m ammals, birds and invertebrates (Gerson and Seaward 1977, Richardson and Young, 1977, Hanley et al. 1989, Rochelle 1980, Maser et al. 1981,1985 1986, Gunther et al. 1983). The largest biom ass of forage and nitrogen

- fixing lichens occurs in late

- successional and old

- growth forests with clean air quality (McCune 1993, Lessica et al. 1991, Neitlich and McCune 1995). Effects of deteriorating air quality on lichens and other sensitive, ecologica lly

critical organisms is an important management concern on national forests and other public lands of the Pacific Northwest (Peterson et al. 1992, Rosentreter 1995, FEMAT 1993, USDA

- FS and USDI

- BLM 1994a and 1994b). The following review describes some of the pollutants affecting lichens in the Pacific Northwest, mechanisms of pollutant accumulation b y lichens, phy siological responses to pollution, methods of study , and provides a summary of regional air quality

studies within the Pacific Northwest utilizi ng lichens. 

## Major Pollutants
A variet y of elements and chemical compounds affecting lichen growth and distribution are found in the atmosphere. Primary pollutants, including sulfur dioxide (SO 2 ), nitrogen dioxide (NO 2 ) and fluoride (F) compounds, remain in the same chemical form after they are emitted into the atmosphere. Secondary pollutants result from chemical reactions involving primary pollutants during atmospheric transport. This class of pollutants includes ozone (0 3 ), perox yacet yl nitrate (PAN), and acid rain components such as sulphuric (H2SO4) and nitric (HNO3) acids. Finally , a third, mixed category of 'air toxics' includes industrial organic compounds, agricultural pesticides, trace metals, and metalloids. Information on effects of compounds in this category on lichens is not extensive (Belnap, et al. 1993). The term "air pollution", as used in this text, designates a mixture of atmospheric smoke, mineral rich dust, sulfur dioxide, nitrogen oxides, sulfur and nitrogen based acidifying compounds, f luorine, photo

oxidants such as ozone and PAN, and air toxics. In general, the air quality in Oregon and Washington is very good compared to other areas of the United States. However, population is expected to grow from 8.1 to 9.4 million people within the next 15 y ears, with concomitant stresses on the region's air resources. For an extensivedescription of current air quality status, emission trends and deposition measurements in the Pacific Northwest region of the U.S., see Eilers et al. (1994). 

### Sulfur d **ioxide** Sulfur dioxide is a by

- product of coal or fuel oil combustion, ore reduction, paper manufacture, many industrial processes, and vehicle exhaust. Natural background concentration is 0.28 to 2.8 mg/m 3 , but values near pollution sources can rise to as high as 200 mg/m 3 or more. Most lichens cannot survive extended periods of SO2 exposure above 60 mg/m 3 (Hale 1981). Sulfur emissions have declined dramatically in the last decade in the Pacific Northwest, attributable to the decline in natural emissions fr om Mt. St. Helens and the closure of the ASARCO copper smelter in Tacoma (Eilers et al. 1994). Total combined emissions of SO2 in 1993 for Oregon and Washington was 211,000 tons, about 15% of the national total. 

### Nitrogen oxide The dominant anthropogenic so urces of NO2 are fossil fuel combustion by stationary sources and vehicles. Although NOx thresholds for lichens have not been established, NOx and other nitrogenous compounds are components of acid rain and photooxidants. Effects of these pollutants on lic hens are discussed in the acid rain section below. Total combined emissions of NOx in 1993 for Oregon and Washington was 571,000 tons, about 2% of the national total. Although NOx emissions in Oregon, outside of Portland, are expected to remain low, emissi ons in Washington are expected to begin increasing again within the next ten ye ars as increases in vehicle use outstrips reductions achieved through emission controls (Eilers et al. 1994). 

### Fluorides Fluorides are released into the atmosphere as by

- products of aluminum, zinc and phosphate ore reduction, but can also occur in fl y ash from coal

- burning installations. Fluoride is a locall y important pollutant around smelters, brickworks, glass making factories and fertilizer plants and are also released in larg e amounts by volcanic eruptions (Richardson 1992). Many lichens are sensitive to this pollutant as it can concentrate in hydrated lichens to more than 200 times ambient concentrations. In general, obvious damage to lichens begins at internal concentrations between 30

- 70 ppm (Gilbert 1971, Perkins et al. 1980). 

### Acid Precipitation Acid precipitation is rain or snow possessing a pH less than 5.6, the pH of pure water in equilibrium with atmospheric concentrations of CO2. Acid rain is produced by regional enric hment of air with SO2 and NO2 which subsequently oxidize to form sulphuric acid (H2SO4) and nitric acid (HNO3). The relative amounts of sulfuric and nitric acids vary in time and geographical location. Deposition actually takes a variety

of forms, includin g dry, aqueous and gaseous deposits of sulfur and nitrogen based acids and oxidant pollutants, including ammonia and its derivatives (e.g. Table 1).Table 1. Known atmospheric sulfates (Dobbins 1979).

**Name Formula**  **Sources Chemical Properties**  **Probable Size Class**  **Diameter (D)**

Sulfuric acid H2SO4 Atmospheric oxidation of SO2; direct from manufacturing Strong acid; very

hygroscopic (drying agent at low relative humidity 0.1

- 1.0 μm

Acid ammonium sulfate NH4HSO4 Oxidation of SO2 with NH3 addition Strong acid;

hygroscopic 0.1

- 1.0 μm
Triammonium acid disulfate (NH4)3H(SO4)2 Oxidation of SO2 plus NH3 Acidic; liquefies at

~65% relative humidity 0.1

- 1.0 μm

Ammonium sulfate (NH4)2SO4 Oxidation of SO2 plus NH3 Weak acid; water

soluble; liquefies at 80% relative humi dity 0.1

- 1.0 μm

Sodium sulfate Na2SO4 Paper pulping b y kraft process Water

- soluble;

liquefies at 84% RH; relativel y inert > 0.5 μm

Calcium sulfate CaSO4 Wind

- blown dust; manufacture of gy psum products Low solubility in water; relatively

inert > 1 μm
Magnesium sulfate MgSO4 Sea spray ; paper pulping Very hygroscopic; relativel y inert and nontoxic > 1 μm
Long range transport processes distribute small particle size, acidifying pollutants to ecosy stems far removed from pollution sources. Both dry and wet deposition are significant routes of acid input to forests in the western United States. Dry deposition is an almost continuous removal of pollutants from the lower atmosphere at ground level. Acidifying compounds commonly deposited as dry

deposition incl ude sulfur dioxide (SO2), sulfuric acid (H2SO4), nitrogen dioxide (NO2) and nitric acid (HNO3). Dry deposition of both nitrate and sulfate in the western U.S. are highest in the summer months (NAPAP 1990). The relative amount of wet deposition declines sea sonally and geographically

with declines in precipitation. In the Pacific Northwest, most of the annual precipitation falls in the winter. Because storms generall y move in from the Pacific, pH values are usuall y above 5. The greatest rainfall acidity

proba bly occurs in the summer, during light summer rains after extended drought . Acid fog and rime ice (frost) are two important t ypes of wet deposition as the ion load is many times the load in rain or snow (NAPAP 1990).Overall annual acid input to forested ecos ys tems in the Pacific Northwest is relativel y low

compared to other parts of the nation (see figures 1 and 2; Eilers, et al., 1994). Although most western forest soils are well buffered (NAPAP 1990), wet and dry de position can accumulate over time. The average concentrations of sulfur and nitrogen species in precipitation throughout the PNW are are not uniform. SO4 2

- and H + ion concentrations in snow along the Cascades increase with increasing latitude suggesting th at precipitation composition in the Washington Cascades may

have been degraded relative to precipitation in the Oregon Cascades (see figures 3 and 4; Eilers, et al., 1994). Poorly buffered alpine

- subalpine sy stems are thought to be most sensitive, although the extent and magnitude of harmful effects in these s ys tems is poorly understood. Figure 1. Precipitation

- weighted concentrations of sulfate (SO 4 2

- ) ions (in mg/L) for the United States for 1992 (Source: NADP/NTN 1993). From Eilers, et al. (1994).

Figure 2. Precipitation

- weighted concentrations of sulfate (NO 3 - ) ions (in mg/L) for the United States for 1992 (Source: NADP/NTN 1993). From Eilers, et al. (1994).Figure 3. Concentration of hydrogen ion (μeq/L) in snow versus latitude (degrees) for 1983 fr om

the Oregon/California border (lat. 42°) to the Washington/Canada border (lat. 49°) as reported b y Laird et al. (1986). The line is a cubic spline fit of the raw data. The rectangles represent samples collected within the national parks. From Eilers, et al. (1994).Figure 4. Concentration of sulfate (μeq /L ) in snow versus latitude (degrees) for 1983 from the Oregon/California border (lat. 42°) to the Washington/Canada border (lat. 49°) as reported b y Laird et al. (1986). The line is a cubic spline fit of the raw data. The dashed line represents the cubic spline fit of nonmarine sulfate. The rectangles represent samples collected within the national parks. From Eilers, et al. (1994).

### Photochemical Toxins

Photochemical toxins are generall y associated with automobile exhaust in cities, and form from the bur ning of fossil fuels. In this group are ozone and peroxy acet yl nitrate (PAN). Ozone is formed b y the breakdown of nitrogen dioxide (NO2) to nitric oxide (NO) and the ox ygen radical (O). The breakdown reaction is cataly zed by hydrocarbons (natural and anthro pogenic) under high sunlight and temperature conditions. The ox ygen radical combines with molecular ox ygen (02) to form the highl y reactive ozone molecule (O3). PAN is formed b y further reaction of ozone with breakdown products of NO2. Photochemical thresh olds for lichens is a controversial topic (see discussion below). In the Pacific Northwest, exposure of vegetation to elevated ozone concentrations occurs downwind of the major metropolitan areas of Seattle, Tacoma, Vancouver, and Portland (NAPAP 1990).
Oz one concentrations higher than 80 ppb occur in the Washington Cascades during the summer months. Federal lands on the western slopes of the Cascades at higher elevations in Washington are particularl y vulnerable, with peak hourl y ozone concentrations rangi ng from 90

- 196 ppb. Although significant effects on forest health have not y et been demonstrated in the Pacific Northwest, based on existing does response data, these ozone concentrations could pose a threat to forest vegetation in Washington (Eilers et al . 1994).Böhm et al (1995) described ozone regimes in and near forests of the Western United States. The

14 EPA calibrated and monitored sites in Oregon and Washington experienced small

- urban, rural

- remote, or remote ozone regimes: Small

- urban ozone regime s occurred in Seattle

- Tacoma, Portland, Eugene, Medford and Salem. Even though Seattle

- Tacoma and Portland are major metropolitan areas, ozone formation potential is lower in the Pacific Northwest (compared to California and east of the Cascades) due to co oler average temperatures, decreased solar radiation intensity and more cloudiness. Small

- urban sites are dominated by urban/small curves, with a few incidences of urban/large and medium curves (See Figure 5). Ozone maxima occur during the summer. Rural

- remote regimes are dominated by remote curves (40

- 90%) with frequent inversion curves. Urban/transport, urban/small curves occur occasionall y (<15%) and urban/large, urban/medium occur rarely (<5%). Pack Forest (south of Tacoma), Columbia County (NW OR), and Clackamas County (NW OR) contained rural

- remote sites. Sites may be impacted b y ozone precursors originating in urban areas although inversion curves also represent downward mixing of ozone from the tropospheric reservoir. Ozone maxima occur during the spring and summer. Remote regimes are dominated by remote curves and included Ol ympic National Park, Port Angeles in NW WA and Crook County in central eastern OR. High ozone concentrations ty

pical of medium and large urban areas do not occur at these sit es. Inversion and small urban curves may occur <15% of the time. Although ozone in the Oly mpic Peninsula has been traced to precursors form Portland and Seattle

- Tacoma, overall these sites are the farthest removed from anthropogenic sources of ozone. Ozone maxima occur in the spring. Figure 5. Seventeen characteristic patterns of hourly ozone concentrations (ppb) identified by Böhm et al. (1995), arranged b y six categories: urban/large, urban/medium, urban/small, urban/transport, inversion, and remote. Ho urly means with standard deviations are plotted. The patterns are labeled using letter and number combinations: the letter refers to the shape of the curve ('A' curves have little diurnal variation; 'E' curves are characterized by larger diurnal fluctuatio ns in ozone), and the number relates to the relative magnitude of the 24

- hour mean ('1' has the lowest mean). From Bohm, McCune, and Vandetta (1995).Part of the difficult y in establishing lichen sensitivity to ozone has been the difficulty in finding

s tudy areas where ozone concentrations are elevated at times of the day when lichens are metabolically

active. Note that all the ozone curves in Figure 5 have maximum values between noon and 6 PM, ty

picall y the time of day

when lichens are least likely to b e active. Some factors which might theoretically increase the likel y hood of ozone damage to lichens would be high elevation (lichens at cooler sites stay metabolically

active longer into the day during the summer time because they

don't dry out as fast), location of study site in an area subject to urban/transport ozone curves (broadens the length of time each day when ozone concentrations are elevated), and location of lichens in microhabitats which keep them moist longer (e.g. among mosses). Potential ar eas for study of lichen impacts from ozone in OR and WA would be the western WA Cascades, especially the NW slope of Mt. Rainier (see Brace 1996), and the Columbia River Gorge National Scenic Area.



### Metals Lichens are supplied with mineral nutrients and heavy metals from precipitation, throughfall, dust, and the underl yi ng substrate from both natural and anthropogenic sources. Natural sources of metals include marine aerosols, leachates from foliage and bark, and suspended particulates derived from local and remote soils and rock. L ichens show varying sensitivity to metals, are good accumulators, and have been used to indicate deposition levels. (See sections on metal accumulation and effects). Some anthropogenic sources for specific metals are listed belo w.Lead (Pb). Pb is the most important heav y metal pollutant in the atmosphere . The main source

of lead in cities is leaded gas; in rural areas or isolated industrial sites, the source is any process involving coal combustion or ore reduction. Nickel (Ni) . Ni is a b y - product of coal combustion, industrial processes, and automobile exhaust. Background levels in the U.S. range from 3 to 9 ppm (Hale 1981). Mercury (Hg). Hg is a minor component of air pollution in the Pacific Northwest from the burning of foss il fuels. Chromium (Cr). Cr is a component of stack fl y - ash and presumabl y travels long distances in the atmosphere; it's presence can be correlated with industrial centers. Zinc (Zn). Zn is emitted from zinc smelters, and is also produced in automobile ex haust (Hale 1981). Titanium (Ti) and Vanadium (V). Ti and V are considered signature elements of petroleum combustion pollution; enhanced levels of these elements strongl y indicate anthropogenic sources (Wetmore 1987b). 

## Mechanisms of Accumulation
Lichens accumulate substances from their environment by a variet y of mechanisms, including particulate trapping, ion exchange, extracellular electroly te sorption, hydroly sis, and intracellular uptake (Nieboer et al.1978). 

### Particulate Trapping
Particles of various elements, such as iron, titanium, aluminum, chromium, and uranium can become embedded in the lichen thallus in the algal and/or fungal lay er under moist or dry conditions (Nieboer et al. 1978, Gough & Erdman 1977). Morphological characteristics that contri bute to particulate trapping include the microtopograph y of the lichen surface (Richardson & Nieboer, 1980), and a large surface area

- to

- volume ratio of the lichen, which can result from a thin thallus, branching, or projections from the thallus such as is idia and phyllocladia (Tomassini et al. 1976, Tomassini 1976). 

### Ion Exchange Ion exchange requires that the lichen must be moist, so that the elements are in ionic form. Uptake of metal ions b y i on exchange is passive and rapid and obey s the principals of m ass and charge balance. Extracellular binding sites, such as at cell walls, are most likely involved. Lichen species vary

in their capacit y for ion

- uptake, as well as the particulate trapping efficiency of their surface morphology . This underlines the impo rtance of correct identification of lichen samples used for monitoring purposes and demonstrates the desirability of collecting the same species for anal ys is from all sampling sites.

### Accumulation of metals

The action of rain, surface water, and passive upw ard diffusion from the substrate likely bring dissolved minerals in contact with lichen thalli (Nieboer et al. 1978, Nieboer & Richardson 1981). The amount of each ty pe of metal ion that can be accumulated by a lichen is dependent upon the uptake characteristics of that particular species and the amount and availability of metal ions in the surrounding environment. Extracellular uptake of metal ions is essentially a passive process of ion exchange determined b y the character of the ligands in the fungal cel l walls. Intracellular uptake is limited by the nature of the metal ion, cell membrane permeability , and the concentration of extracellular ligands with affinity for cations (Tyler 1989). Lead. Lichens are very

efficient accumulators of Pb, through aerosol s, particulate metal fallout, or acid rain. Pb is bound to insoluble anionic sites, accumulated extracellularly , and concentrated in the medulla. Once bound, Pb is not easily removed by

rain or wind. The toxic effects of Pb on lichens is minimal. Some spec ies can accumulate up to about 2000 ppm, after which concentrations do not increase, indicating a degree of phy siological turnover. Survey s of Pb content have proved useful in establishing background (preindustrial) values which can be referenced to present levels (Lawrey & Hale 1981). Nickel. Ni is taken up b y algal cells in lichens. Mercury . Hg is readil y accumulated b y lichens. Background values were measured at approx. 0.223 ppm, while specimens 8 km from a chlor

- alkali plant contained 0.53 ppm (L odeniu s, 1981). Chromium. Schutte (1977) reported Cr accumulation in lichens of Ohio increased two to ten

- fold over turn of the century

concentrations. Copper. Nieboer et al. (1979) documented two phases of Cu uptake in Umbillicaria muhlenbergii . In the first ph ase copper ions bond to receptor sites on algal cells. In the second phase, copper ions bind to the fungal cells, as evidenced b y K+ efflux. Zinc. Zn is found in fairly large concentrations (200

- 600 ppm) in lichens near zinc smelters. 

## Physiological **Responses to Pollution Lichens as pollution monitors**  Lichens and bryophy

tes possess a number of characteristics that make them suitable biomonitors for air pollution. Man y lichen species have large geographical ranges, allowing study of pollution gradients over long distances. L ichen morphology does not vary

with the seasons, and accumulation of pollutants can occur throughout the y ear. Li chens are u sually very long lived. Lichens are also suitable biomonitors due to their sensitivity to pollution. Several factors contribute to this sensitivity . Water and gas are exchanged over the entire lichen thallus. Because they lack roots, lichens do not have ac cess to soil nutrient pools and must depend on deposition,water seeping over substrate surfaces, atmospheric and other comparativel y dilute sources of

nutrients. Thus, their tissue content largely reflects atmospheric sources of nutrients and contaminants . L ichens also lack the protective tissues or cell t ypes necessary to maintain a constant internal water content. Many lichens pass through multiple wetting and drying cy cles during a day . When hydrated, nutrients and contaminants are absorbed over the ent ire surface of the lichen. During deh ydration, nutrients and many contaminants become greatl y concentrated in the lichen tissues, i.e., absorbed to cell walls, cloistered inside organelles or cry stallized between cells. Different contaminants have different release rates depending on the way they are stored. Nutrients tend to be more easily remobilized than metal contaminants. During heavy rains, nutrients and pollutants are gradually leached. A dy

namic equilibrium thus exists between atmospheric nutrient/p ollutant accumulation and loss, which makes lichen anal ys is a sensitive tool for the detection of changes in air qualit y. Cryptogams have consistently shown higher metal levels than higher plants (Rühling& Tyler 1968, Huckaby 1993, LeBlanc et al. 1974, Hoc king et al. 1978, Solberg & Selmer

- Olsen 1978, Arafat & Glooschenko 1982). In one study , lichen metal levels reflected the prevailing wind direction to a greater extent than did pine needles (Laaksovirta& Olkkonen 1977, 1979). Different morphological ty pes of lichens exhibit differing levels of sensitivity to pollution as a result of combined factors that are not well understood. In general sensitivity increases in the following series: crustose < foliose < fructicose. 

### Effects of Pollution on Lichens A my ri ad of pollution effects on lichens have been described. At the level of the whole plant, investigators have described decreases in thallus size and fertility (DeWit 1983, Kauppi 1983, Sigal & Nash 1983), bleaching and convolution of the thallus (Sigal & Na sh 1983), restriction of lichen occurrence to the base of vegetation (Sigal & Nash 1983, Neel 1988), and mortality of sensitive species (DeWit 1983, Denison & Carpenter 1973). Microscopic and molecular effects described include reduction in the number of a lg al cells in the thallus (Holopainen 1984), ultrastructural changes of the thallus (Anderson & St. Clair 1983, Hale 1983, Holopainen 1984, Pearson 1985), degradation of photosy nthetic pigments (Kauppi 1980, Garty et al. 1993), altered photosy nthesis and r espiration rates (Sanz et al. 1992, Rosentreter & Ahmadjian 1977), and elevation in the content of heav y metals in the thallus (Addison & Puckett 1980, Carlberg et al. 1983, Gailey

& Llo yd 1986 a, 1986b and 1986c; Gough& Erdman 1977, Lawry 1986). Huebert e t al. (1985) concluded that peak exposures may be of primary

importance in determining the survival of lichens in industrial areas. 

### Sulphur dioxide SO2 is considered to be the primary factor causing the death of lichens in most urban and industrial areas, with fruticose lichens being more susceptible to SO2 than many foliose and crustose species (Seaward 1987). The observed effects of SO2 include decreases in respiration and photosy

nthesis, increases in membrane permeability , increases in K+ influx and ions lost, and ultrastructural changes (Belnap et al. 1993). Damage to the algal component of the thallus is evidenced b y its discoloration. The entire thallus dies soon after algal cells are damaged (Hale1981, Wetmore 1985). Low pH increases the toxicity of SO2 action (Farmer et al. 1992)). Nieboer et al. (1979) determined that copper also enhances the effects of SO2. There are three general mechanisms by which SO2 exerts it's toxic effects upon the lichen thallus. At low pollution levels, enzy me s ys tems with in the thallus may become activated. This can be observed as increases in glucose

- 6

- phosphate dehydrogenase activity, or an increase in the levels of glutathione and total protein SH. Heightened enzy me activity

can be used as a bioindicator for the non

- vis ible injury metabolism and detoxification of absorbed SO2 (Rabe and Kreeb 1980, Miszalski& Zeigler, 1979, Grill et al. 1980). At higher levels, SO2 deactivates enzy mes b y chemical modification (sulfitoly sis) leading to reduced metabolic activity and loss o f membrane integrit y (Zeigler 1975 and 1977, Nieboer et al. 1976). Italso binds to the central metal atoms of enzy

mes, adversel y affecting membrane function and cell osmolality (Malhotra and Khan 1980, Gunnison et al. 1981, Mansfield and Freer

- Smith). In addition, SO2 competitively

inhibits carbonate (HCO3) and phosphate (H2PO4) interactions with enzy mes (Alscher

- Herman 1982). Table 2. General Mechanisms of SO2 toxicity
**Type of reaction Observed/expected response or**  **injury References**  Enzy me deactivation: (e.g. sulfitoly sis)

- chemical modification

- binding to metal centers (Vit B12; Fe)

- inhibition Reduced metabolic activity ; loss of

membrane integrit y, membrane function and cell osmolality , (competition with HCO3 - , H2PO4 - ) Zeigler 1975, 1977, Nieboer et al. 1976, Malhotra and Khan 1980,Gunnison et al. 1981, Mansfield & Freer

- Smith 1981, Alscher

- Herman 1982 Stimulation of enzy

me sy stems: (often in response to low pollution levels)

- Glucose

- 6

- phosphate

dehydrogenase,

- Increas es in glutathione and total protein SH Use of increases in enzyme activity

as a bioindicator for non

- visible injury and detoxification of metabolism absorbed SO2 1979, 19 Rabe & Kreeb, 80, Miszalski & Zeigler, 1979, Grill et al. 1980

Reaction with reactiv e biomolecules:

- chemical (bisulfite adducts)

- redox (acts as electron acceptor/donor at pH 7) Modification of metabolic precursors

and products; interference with electron flow in photos ynthetic and respiratory

electron

- transport chains. Puckett et al. 19 73; Ziegler, 1975; 1977; Gunnison, 1981.Nieboer et al, 1976; Shapiro, Petering & Shih, 1975; (108)


### Effects of Acid RainNatural and anthropogenic variations of substrate and water acidity are a major factor governing the composition and health of bry ophy te and lichen communities (Farmer, et al. 1992). At the level of the individual thallus, acidity affects nitrogen fixation , photosy nthesis, growth, and cellular ultrastructure (Belnap et al. 1993). Besides these direct effects on phy siological processes, environmental acidity can affect cry

ptogamic communities indirectly by altering substrate chemistry , thereby affecting spec ies diversity and composition at the community level. **Bark pH interactions** Ecological studies of epiphy te communities have shown that, even in unpolluted areas, natural variations in bark chemistry have a major impact on which species are present ( Farmer, et al. 1992). For example, Hale (1955) found that species of Physcia , Xanthoria , and Candelaria were more common on trees with bark pH values of 5.3 to 6.6. Many factors can influence bark pH. Tree species differ in this characteristic; for instance, conifers generall y have lower bark pH than angiosperms. Other influential facto rs affecting the number and t ype of epiphy tes present include the buffering capacity

of the bark, dust accumulation, and the age of tree. Individuals of a single tree species can have markedly different bark chemistries and epiphy te communities (Farmer et al. 1992). For example, Gilbert (1970) showed that the pH of Fraxinus could vary on a single tree from 5.7 along a nutrient streak to 2.2 on the underside of a branch. Gough (1975) speculated that continuous peeling off of bark scales may be one reason for little epiphy tic cover on Pinus ponderosa in Colorado. He found that P. ponderosa bark loses water faster than the bark of aspens and Pseudotsuga menziesii, Abies lasiocarpa, Picea engelmanii, and Pinus contorta . Distribution of lichens on trees may also be related to pH. The pH of the outermost newly

- exposed orange bark of trees over 4 dm in diameter tended to be significantl y lower (pH 3.5

- 4.27) than that of bark in furrows (pH 3.71

- 4.66) of the same trees and bark of smaller trees (pH 3.76

- 4.84). The pH of bark on twigs averaged about 4.78, significantl y higher than bark on trunks; lichen cover on branches and twigs was generally more continuous on than on tree trunks. The presence of pollutants also affects bark pH. Gilbert (1965) studied the pH of the bark of Fraxinus excelsior and Acer pseudoplatanus in and around Newcastle. While the average pH value of F. excelsior was 4.2 both at 31 km from Newcastle and in the cit y itself, the bark pH of A. pseudoplatanus dropped from 4.8 at 31 km from Newcastle to 3.5 in Newcastle. Robinia psuedoacacia trees growing in the vicinit y of a pulp mill (Hoffman 1974) were found to have bark pH values of 6.4 to 5.2, with the lower values generall y corresponding to the trees closest to the mill. British researchers have no ted that bark pH is a critical factor in the establishment of acidophilous species of the Parmelia laevigatae community , normally associated with high elevation woodlands. With continuing acidification these species can replace the lower elevation Lobaria community . Acid rain can reduce pH of both tree bark and lichen thalli (Farmer et al., 1992). **Soil chemistry interactions**Variations in soil chemistry may be important. For example, Norwegian oaks with Lobarion species grew mainly

on nutrient rich soils. V ariations in bark chemistry may be affected by variations in soil cation availability . The relationship between soil and bark chemistry is further complicated b y the ability of trees with acidic stemflow to acidify the surrounding soil, as demonstrated for Fagus sylvatica in Sweden (Farmer et al, 1992). Bark and soil acidification is caused b y the total H + ion input to the sy stem. The critical H+ ion deposition load for epiphy tes is the highest load that will cause a change in surface bark chemistry , particularly

of base saturation, that can be countered by natural chemical and weathering processes (Farmer et al. 1992). For epiphy tes, extreme events are evened out by the buffering capacity

of bark, so that the best predictions of epiphy te response are probabl y made on annual averages of pollutant loading. L ichens may respond to short episodic pollution event s within a local area that may be undetected by

conventional anal ys is.
**Other substrate interactions** Evidence from field observations, intensive studies of the environment at contrasting sites, and field and lab experiments (Farmer et al. 1992) show that ac id rain can affect lichens both directly and indirectl y b y the acidification of substrata. Seaward (1987) writes, "Regional pollution over wide areas has reduced lichen diversit y, and has favored the expansion of a relativel y small number of species formerly having narrower ecological requirements and/or more restricted distributions such as Buellia punctata, Lecanora conizaeoides, Lecanora muralis, Parmelia incurva, Parmeliopsis ambigua, Phaeophyscia orbicularis, Scolisciosporum chlorococcum, Xanthoria can delaris, and Xanthoria elegans ." Pollution most greatly affects lichen species sensitive to high concentrations of H + and Al3 + . In addition, when pollution occurs in poorl y buffered environments, species such as Lobaria which normally thrive on mildly

acid ic to neutral substrate, will also be negativel y affected (Farmer et al. 1992). **Acid fog** Fog, which often increases with increasing elevation, contains higher levels of dissolved ions, including H + , than precipitation rain or snow. Although the effects of occult deposition from fog on epiphy te communities have not y et been critically examined, acid fog has been implicated in changes to terricolous cryptogamic floras (Wolseley & James 1992). Desert lichens, which obtain most of their water from fog and dew, are particularl y vulnerable to air quality and weather pattern changes (Follman 1995, Nash 1996). **Sensitivity of lichens to sulfur dioxide vs. acid rain** Species tolerant to acid rain need not be the same as those tolerant of SO2. Usnea species intolerant o f SO2 grew on bark of low pH, while Parmelia sulcata was tolerant of SO2, but not of acid rain (Gilbert 1986). **Nitrogen enriched acid rain**The nitrogenous component of acid rain can produce a fertilizer effect on lichens and cause

floristic changes. Søchti ng (1990) survey ed the tissue nitrogen content of reindeer lichens in Denmark. In unpolluted areas, tissue levels were 0.26

- 0.49 % while in areas of wet deposited acidity , values were 0.70

- 0.73 % and visible injury could be found. However, the problem of defining critical loads for nitrogen has not been resolved. Presently , this is determined by the biological response of the sy stem, and is very hard to quantify , compared to H+, which is measured b y buffering capacity (Famer et al. 1992).
**Lichen structure and physiology**  The effects of acid rain on lichen structure and physiology include a decrease in nitrogen fixation, photosy

nthesis, and growth, and changes in cellular ultrastructure (Belnap et al. 1993). Hutchinson et al. (1986) treated Cladina rangiferin a and Cladina stellaris with artificial acid rain for more than five ye ars. They found reduced photosy

nthesis and chlorophyll levels, and decreased dry

weight and podetial height of thalli of both species receiving pH 2.5. No effects were observed at pH 3 and above. Nitric acid applied alone at pH 2.8 caused dry weight increases of 62% over controls in C. rangiferina with an increase in tissue nitrogen levels. Sulphuric acid applied alone caused a significant reduction in dry weight. 

### Effects of Fluorine
The ability

of lichens to accumulate F is a function of relative humidity , which determines the moisture conditions of the thallus. In general, obvious damage to lichens begins at levels of 50

- 70 ppm (Gilbert 1971). The effects of fluorine include a decrease in respiration and photosy nthesis, an increase in membrane and thallus permeability with a concomitant loss of ions, and changes in cellular ultrastructure (Belnap et al. 1993, LeBlanc et al. 1971). Lichens exposed to ambient F at 4 mg F/m 3 accumulated F w ithin their thalli, and eventuall y surpassed the critical concentration of 30

- 80 ppm (Perkins et al., 1980). Both in the field and in the lab, whenever the level of F within the thallus exceeded 80 ppm, chlorosis was observed. Subsequently , all the photosy nthetic pigments were degraded and the lichen thalli disintegrated. In the field, wherever chlorotic transplants were found, high levels of F on lime filter papers were also found, suggesting that ambient F was the cause of the lichen injury . The lichens u sed in this study

were Cladonia cristellata, C. polycarpoides and Parmelia plittii (Nash 1971). Perkins et al. (1980) showed that F accumulation by corticolous lichens was negatively correlated with distance from an aluminum reduction plant and that lichen damage and reduced abundance was due to F. SO2 emitted in combination with HF from a mix of industries in Whatcom County , Washington, was associated with a serious depletion of the lichen flora, even though emission levels were within acceptable limits ba sed on human health standards set b y the US Environmental Protection Agency (Taylor & Bell 1983). 

### Effects of photochemical toxins Ozone, PAN and nitrogen oxides are toxic to lichens in sufficient concentrations. Ozone effects noted by researches have included decreases in photosy nthesis, reduction in the geographicaldistribution of a species, and morphological and ultrastructural ch anges. PAN can also affect

photosy nthesis and cause ultrastuctural changes, while NOx effects are expressed in the loss of chloroph yl l pigment (Belnap et al. 1993). In higher plants, oxidant air pollution is thought to disrupt normal pathway s of energy by altering cell membranes; no comparable data for lichens exists, and experiments to determine the effects of ozone on lichens have yielded contradictory results. Field studies in the L os Angeles basin (Neel 1988, Sigal & Taylor 1979, Sigal & Nash 1983) show a broad correlation between lichen loss and increases in phototoxins in the past 80 years; however, phototoxins are difficult to research, as they are produced over a large area in low concentrations and have at most long

- term or chronic effects on lichen s (Hale 1981). Ozone appears to have little visible effect on roadside lichens (Lawry & Hale 1979). L ichen communities in the Indianapolis area were insensitive to differences in ozone levels. This could be due to an intrinsic species insensitivity to ozon e, or from the fairl y narrow ozone dose range observed. Over the three year stud y period the means of ozone deposition differed b y 12 mg/m 3 across sites, whereas means for SO2 differed b y 17 mg/m 3 (McCune 1988). Photochemical toxins may

have s yn ergistic ef

fects when combined with other pollutants, especially under low pH conditions but this sy nergy has not been conclusively documented (Farmer et al. 1992). Ruoss and Vonarburg (1995) studied lichen metabolic activity in relation to ozone concentrations at three elevations in central Switzerland. They also fumigated lichens at 300 ppm for 34 day s. While the fumigated lichens showed damage (bleaching and discoloration), the specimens from the field experiment were not differentially affected by

ozone. The autho rs found that ozone levels never exceeded 120 ppm when relative humidity was high enough to sustain lichen metabolism and, therefore, lichens were not susceptible to ozone during periods of high ozone concentrations. While lichen diversity and abundance in creased with elevation the highest ozone levels occurred at the highest elevations. The authors felt lichen diversit y and abundance was primarily impacted winter pollution, i.e. acidic deposition, which decreased with elevation. 

### Effects of metals
Metal ion s can be classified into three groups. Class A ions: K+ , Ca2+ , and Sr2+ are characterized by a strong preference for O2

- containing binding sites and are not toxic. I ons in the B class: Ag + , Hg + , Cu+ , tend to bind with N and S

- containing molecules, and are extremely toxic to lichens even at low levels. Borderline class ions: Zn 2+ , Ni 2+ , Cu 2+ , Pb 2+ , are intermediate to Class A and B ions. Borderline ions, especially those with class B properties, e.g. Pb 2+ , Cu 2+ , may be both detrimental by

themselves and in combination with sulphur dioxide (Nieboer & Richardson 1981). The toxicity

of metal ions is the result of three main mechanisms: the blocking of essential biological functional groups, the displacement of essential metal ions, and the modification of the a ctive conformation of biomolecules. In studies of specific ions, Nieboer et al. (1979) found that Pb 2+ also affects cell wall permeabilit y and causes K+ efflux, which is correlated with a decrease in carbon fixation.Limited field data is available concern ing threshold concentration of metals in lichens. Tolerance

mechanisms mainly

involve immobilization of the toxic metal ions in biologically inactive forms. Tolerance to metals may be phenoty

picall y acquired, but sensitivity of lichens to elevated tissue c oncentrations of metals varies greatl y between species, populations, and elements (Tyler 1989). 

## Methods of Study Lichens have been used to assess deposition and air quality in hundreds of studies worldwide. Reviews of the literature on this topic include R ichardson (1992), Stolte et al. (1993), Nash & Wirth (1988), Nash (1989) or see the bibliography to this document. The Lichenologist publishes an ongoing series "Literature on air pollution and lichens". Below we provide examples of different approaches. 

### M **ethods using naturally occurring lichens Distribution mapping**  Distribution mapping is the classic field method for studying lichens. It was also the first method used to indicate air quality with lichens (Showman 1988). Different kinds of maps can record the total number of species per site (richness), the percentage of the quadrants in which particular species can be found (frequency), presence or absence of indicator species, and the estimated or measured cover. Mapping lends itself well to longitudinal studies that document temporal change. Using this method, Sigal and Nash (1983) compared lichen communities in the San Bernadino and San Gabriel Mountains of California in 1983 to epiphytic macrolichens on conifers reported b y Hasse in the same area in 191 3. A number of species were no longer found: Nodobryoria abbreviata, Bryoria fremontii, Nodobryoria oregana, Alectoria sarmentosa, Calicium viride, Cetraria canadensis, Platismatia glauca, Evernia prunastri . Other species, such as Cetraria merrilli, Parmel ia quercina, Ramalina farinacea , and Cladonia spp., were found in only trace amounts. Species that were present included Hypogymnia enteromorpha, Melanelia elegantula, Melanelia subolivacea, and Letharia vulpina

. One disadvantage of presence/absence distri bution mapping is that a species must disappear before an effect can be registered. To address this shortcoming, De Sloover & Leblanc (1968) developed the Index of Atmospheric Purity (IAP). At each site, this method assigns numerical values to characterist ics such as frequency and cover. Then, using a mathematical formula, it reduces these values to a single IAP value for each site. The IAP values are then mapped. The primary limitation of IAP values is that they are not comparable from one geographic regio n to another. Even in the absence of pollution, climatic differences between regions can produce changes in the lichen communities that underlie the IAP values (see Hoffman 1974).
**Gradient studies**Gradient studies correlate available pollution exposure dat a with observations of visible injury ,

species richness and species abundance. Linear regression anal ys is is used to investigate the interrelationship between variables, and to predict one variable when the other is known. Taylor and Bell (1983) used regre ssion to show that, at the 95% confidence level, there was a functional relationship between distance from a smelter in Tacoma, Washington and total lichen cover, with an increase in cover with increasing distance from the smelter. Compared to data collect ed in 1976, the SO2 emissions from the smelter and the mean concentrations of sulfur found in leaves of alder trees at the study sites had both decreased. This suggested that the correlation between lichen growth and distance from the smelter was more or l ess established during earlier periods of greater ambient SO2 levels. Multivariate anal ys is of epiphy tic macrolichen survey

data was used to assess air quality in the southeastern United States (McCune et al. 1994). Two major gradients were revealed. The f irst gradient corresponded to a climatic gradient from the coast to the Appalachian mountains. The second gradient corresponded primaril y to air quality, with pollution

- tolerant species and lower species richness at one end of the gradient and pollution

- se nsitive species and high species richness at the other end of the gradient. The gradient model was then used to calculate air pollution impact scores for sites to answer questions regarding the spatial pattern and trends of air quality

effects on forest re sources. 

### Laboratory Measurements Laboratory

methods used to study

the chemical, morphological and ph ys iological responses of lichens to pollution include elemental anal ys is, phy siological measurements, transplant studies, fumigation studies and photograph y . **Elemental Analysis** Methods of elemental analy sis include atomic absorption spectrophotometry , neutron activation anal ys is, and X

- ray

fluorescence spectrometry (Nieboer & Richardson 1981). Commonly anal yz ed elements include: S, N, Zn, F, Pb, Ni, Cu, Fe, Mg, Mn, Cr, Cd, Zn, Ca, and several rare earth elements. A large body of information is developing on elemental content of lichens, both in natural states and under polluti on stress (Nash 1996). Hale (1981) recommended that every baseline study

should include elemental analy sis of lichen indicator species for future reference. Using the gradient method, element concentrations within the lichen are usually observed to increase as the distance to the suspected source decreases. Gough and Erdman (1977) used linear regression to evaluate the relationship between distance from a coal fired power plant and metal levels in Xanthoparmelia chlorochroa . However, as Puckett (1988) point out, concentrations of many elements will not reach zero at large distances from pollution sources because they have essential nutritional roles or are norm al components of the lichen when growing in its natural environment. When, both natural and anthropogenic sources contribute to the metals measured, identifying the relative contribution from each source becomes a complex task. Puckett (1988) reported a me thod of calculating "enrichment factors" to compare the concentration of metal within a plant with potential sources in the environment. The equation isEF = x/reference element in lichen / x/reference element in crustal rock

Stable sulfur isotope ratio s in combination with multi

- element anal ys is of lichens were used to examine the influence of emissions from two coal

- fired power plants i n the Yampa Valley on pollutant deposition in the Mt. Zirkel Wilderness of northern Colorado (Jackson et al., 1996). Coal

- fired power plants ty pically emit SO2 with a stable isotope ratio 34 S/ 32 S characteristic of the coal combusted. Stable S isotope ratios in Usnea were significantly heavier (more positive) in the wilderness and nearby sites compared to more distant reg ional sites and corresponded well with sulfur isotope ratio found in snow in the same area and average isotope ratios in coal used by the power plants. Sulfur isotope studies are most easily interpreted when the point source of concern is the predominant s ource of sulfur in the study area. Xanthoparmelia cumberlandia from the Mt. Zirkel area were also elevated in sulfur, nitrogen, potassium, sodium and phosphorus compared to more distant regional sites (>100 km). Determining relevant elements is an important part of any multi

- element study . In Switzerland, Herzig et al. (1989 & 1990) used multivariate analy sis to compare element concentrations in Hypogymnia physodes to total air pollution as assessed by lichen communities using the IAP index and an instrumented monitoring network. Four groups were discerned.

 Group 1: Ca. Calcium was the only element which increased with improving air quality

.  Group 2: Pb, Fe, Cu, Cr, S, Zn and P. Concentrations of these elements decreased in distinct curvilinear gradients with decreasing total air pollution. For example, the concentration of Pb was reduced six

- fold in the "very low pollution" compared to the "critical air pollution" zone. These elements were strongl y correlated with annual average atmospheric deposition measur ements detected by the instrument network.

 Group 3: L i, Cd, Co. Concentrations of these elements were lower in the "very low pollution" zone than in the "critical air pollution", but the gradients were not strictly curvilinear.

 Group 4: Al, B, K, Mg, Mo, N a, Ni, Sn, Cl and organic S. These elements showed no clear gradients to the total air pollution.. Other cry ptogams, especially

mosses, have also been successfully used as pollution monitors (Winner 1988). Between 1989 and 1992, twent y - one European countri es participated in an unprecedented effort to map atmospheric toxic metal deposition based on moss analy sis (Rühling 1994). Colored contour maps of the continent for As, Cd, Cr, Cu, Fe, Pb, Ni, Vd, Zn deposition were made from metal analy ses of Hylocomium splendens , Pleurozium schreberi and some secondary species. A strong linear correlation between average annual atmospheric deposition and concentration b y moss was observed for As, Cd, Cr, V, and Zn. which could be used to calculate atmospheric deposition. Innorthern Europe, metal concentrations generall y increased from north to south with local toxic metal enhancements superimposed on the regional background pattern at the large smelting region of Russia in the western Kola Peninsula. In central Europe, m ajor enhancements were detected around the industrial regions of the Ruhr area in German y, western Czech Republic and adjacent eastern German y, and southern Poland. **Physiological responses**The phy siological responses of lichens to pollution can be observed and measured directl y.

Plasmoly sis of cells of the algal component of the thallus can be measured microscopicall y. Respiration rates have also been measured. Von Arb et al. (1990), observe d changes in the fine structure of chloroplasts. Chlorophyll content, as a measure of SO2 damage, was determined spectroscopically

or by microfluorometry

by Beltman et al. (1980), Kauppi (1980), Kauppi and Mikkonen (1980) and Gries et al (1995). Transplant s Lichen transplants are used to assess air quality in areas where lichens are absent or sparse. Richardson (1992) reviews the use of transplants to assess air quality in urban environments and to monitor contaminants in air and water. Pearson (1993) discu sses advantages and limitations of transplant methods. Ideally , healthy lichens are transferred from an area where they occur naturally to an area to a test area. Changes in phy siology or element accumulation as a result of exposure to pollution are then s tudied. Phy siological studies are most likely to be successful when using species well within their normal range of adaptation both at the source and at the test area. Transplant techniques for common lichens in Pacific Northwest environments west of the C ascade Range crest are discussed by Shirazi et al. (1996) McCune et al. (1996) and Sillett (1994). Hypogymnia physodes , a fairl y pollution tolerant lichen widel y distributed in the Pacific Northwest, has been one of the most frequentl y transplanted lichens world

- wide (Brodo 1966, Holopainen 1984, Farkas et al. 1985, Gailey et al. 1985, Vestergarrd et al, 1986, Søchting 1995).

**Fu migation**  Fumigation studies involve the controlled introduction of pollutants to lichens in either a closed chamber that allows f

or air movement, or in a chamber

- free field environment. The purpose of this kind of study

is to develop quantitative relationships between concentrations of air pollutants and various anatomical and phy siological lichen responses. The anticipated outcome of air pollution fumigation experiments is a clear exhibition of dose

- response relationships. Thus, fumigations for a long period of time at a low pollutant level would be expected to have the same effect as a shorter fumigation at a higher concentration where the dose, measured as concentration multiplied by time, is equal for each fumigation (Sanz et al. 1992) . The most commonly

measured responses are phy siological, such as photosy nthesis, respiration, nitrogenase activity

in blue green ph yc obionts, K+ e fflux and/or total electroly te leakage from the thallus, and pigment status. A problem inherent in fumigation studies is that different portions of the thallus differ in their sensitivity

and phy siological response to pollutant exposure. Factors such as th e hydration level and phy siological status of the thallus, as well as the differential reactions of the algal component vs. the fungal component, can affect its sensitivity to exposure. In addition, the concentration and duration of exposure, and the envir onmental conditions under which exposure occurs, can also affect sensitivity

(Fields, 1988).A strength and limitation of fumigation studies has been the tendency of researchers to expose their experimental material to only a single pollutant under a given set of conditions. Air pollution involves multiple contaminants (Belnap et al. 1993), and it is important to develop methods to test for the synergistic effects of pollutants found in the field (Fields, 1988). Farmer et al. (1992) also stressed the need f or long term fumigation studies employing realistic pollutant concentrations, as the way to demonstrate conclusively whether the levels of pollutants measured in the field were actuall y damaging to cry ptogamic flora. Researchers evaluating fumigation experiments also need to remain mindful of the ecological conditions under which the studies are carried out (Farmer et al. 1992). For example, water relations play

a significant role in the uptake of SO2 by lichens (Grace et al. 1985). Dry lichens absorb far l ess SO2 than moist ones and are resistant to SO2 fumigation (Coxson 1988). The low humidities used in some studies (Showman 1972, Beekley & Hoffman 1981), could have completely

dried the thalli. Thus, the finding that responses in some studies appear to be independent of exposure time is not surprising (Fields & St. Clair 1984; Huebert et al. 1985). With poikilohydric plants such as lichens, it is difficult to maintain the water status and hence metabolic activity

of a sample for more than a few hours. Comp lex laboratory equipment is necessary to successfully conduct fumigation studies. The relationship between seasonal environmental conditions and pollution is an important adjunct to the findings of SO2 fumigation studies. For instance, NO2 reaches high lev els in the winter when lichens and bry ophy tes are more likely

to be wet and phy siologically active. O3 reaches its peaks in the summer. However O3, unlike SO2 and NO2 is not readil y soluble in water and may

damage cell membranes even in a dry state (Farmer et al. 1992). The time of ye ar also affects lichen net assimilation rates; for example the net assimilation rates for Peltigera praetextata and Parmelia sulcata in June increased five times over the net assimilation rates recorded from October to March (K ershaw 1972). Even though the various gaseous pollutants have different ph ys iochemical properties, they affect the phy siological s ys tems of lichens in similar way s (Fields 1988). In general, greater concentrations of, and longer intervals of exposure to SO2 result in re duced photosy nthesis and reduced respiration, although at low levels and short exposures SO2 can enhance 14 C fixation. SO2 also produces membrane damage leading to K+ efflux and total electroly te leakage in a time dependent manner, with massive K+ release occurring at increasingl y lower SO2 levels as exposure time is increased (Fields 1988). Oxidants, O3 and PAN reduce photosy nthesis. The order of sensitivity

of lichen phy siological processes to fumigation appears to be: N2 fixation > K+ efflux/total electr ol yt e leakage> photosy nthesis and respiration > pigment status. Table 4. The ph ys iological responses of lichens after exposure to fumigated pollutants (Farmer et al., 1992, Fields 1988).

**Response Pollutant**  Reduced photos ynthesis SO2, NaHSO3, Na2S2O5, O3, PAN. Reduced respiration SO2Decreased chloroph yl l content SO2, NO2, HF

In creased electrol yt e leakage, loss of K SO2, HF? Reduced nitrogen fixation SO2, H2SO4 > loss w/< pH > exposure, NaHSO3, NaF **Photography** Some researchers have used photograph y as a way of monitoring air quality (e.g. Hale 1981). Marked sites can be rephotographed over time to compare growth rates and condition of individual lichen colonies. Care must be taken to distinguish between air poll ution and local habitat stresses. 

## Air quality monitoring studies using lichens in the Pacific **Northwest** Air pollution studies using lichens in the Pacific Northwest west of the Cascade ranges are graduall y accumulating in number. Rhoades (1988) examined lichen communities and chemistry in Oly mpic National Park. (Gough et al. 1988) established baselines in Re dwood National Park in northern California. Taylor and Bell (1983), and Johnson (1979) examined lichens in the Seattle

- Tacoma area. Moser et al. (1983) investigated the impact of emissions from Mt. St. Helens on two lichen species of South

- Central Washingt on. Geiser et al. (1994) working in southeastern Alaska, established element concentration baselines for Cladina rangiferina, Lobaria oregana, Hypogymnia enteromorpha , and Alectoria sarmentosa , lichens widespread in the Pacific Northwest. Rhoades (1996) re viewed the literature of tissue chemistry for lichens occurring in the Pacific Northwest. McCune and Geiser (1997) proposed sensitivity ratings for ninety

- three lichens using field based, lichen community studies in western Washington and Oregon. Dennison and Carpenter (1973) documented air pollution in the Willamette Valley of Oregon. The USDA Forest Service and USDI National Park Service established air quality monitoring methods for using lichens (Stolte et al. 1993) and the EPA/USFS sponsored national F HM program has protocols (McCune 1994 ) currently used in adapted form (Geiser 1995) within the Pacific Northwest region of the USDA Forest Service for lichen monitoring. This program has accumulated a large database of lichen survey and chemistry

informat ion on a 3.4 mile grid across five national forests and the Columbia River Gorge National Scenic Area. East of the Cascade Crest, published studies of air quality assessments using lichens are scarce. However, due to similarities in climate, vegetation, an d lichen flora, those interested in Oregon and Washington air quality issues may

find it useful to refer to some of the various baseline and point source studies in I daho and Montana (Rope and Pearson 1990, Hoffman 1974, St. Clair and Newberry

1995), W yomi ng (Gough and Erdman 1977) and Colorado (Jackson et al. 1996, Hale 1981).