Impacts and Effects Indicators of Atmospheric Deposition of Major Pollutants to Various Ecosystems-A Review

In this paper, we review the current understanding on ecosystem and human health impacts from the atmospheric deposition of acidifying pollutants, eutrophying nitrogen (N), polycyclic aromatic hydrocarbons (PAHs), mercury (Hg), trace metals, and ozone (O3), as well as the biological indicators that have been used to assess the health of ecosystems following exposure to these pollutants. We provide overviews of the impacts of deposition for these pollutants and discuss the currently known biomonitors for each pollutant. The deposition of acidifying pollutants impacts terrestrial ecosystems by altering plant physiology and growth and by increasing plant susceptibility to stresses that can be indirectly damaging to the health of fish and birds. Indicators of the deposition of acidifying pollutants include soil base cation content and acid neutralizing capacity, among others. Eutrophying N deposition has been studied extensively; N enrichment directly impacts vegetative plant species cover, richness, growth rates, and susceptibility to other stressors. It indirectly impacts wildlife through changes in their habitats and food sources. Indicators for N deposition include changes in plant species and in tissue and litter N content. The deposition of PAHs has been found to cause significant damage to plant organisms and to be carcinogenic and mutagenic to humans and animals. Useful biomonitors of PAH deposition include lichens, mosses, and pine needles. Deposited Hg can undergo methylation (in the presence of sulphur reducing bacteria); bioaccumulation of methylmercury is highly toxic to animals. Effective biomonitors of Hg contamination of aquatic ecosystems are fish and marine birds. The impacts of O3 are well understood, with well-established “flux” models being vast improvements on the previous AOT40 approaches. This review highlights the impacts that the above-mentioned pollutants have on terrestrial and aquatic organisms and the biomonitors that are currently being used to assess the deposition levels and effects of these pollutants.


INTRODUCTION
Atmospheric deposition is a major source of pollutants to terrestrial (e.g., forests, grasslands, ombrotrophic wetlands) and aquatic (rivers, lakes, oceans) environments (Lepori and Keck, 2012).Other non-depositional processes (e.g., effluent discharge, run-off, leaching) are also important pathways for pollutant input, especially to aquatic ecosystems (Vodyanitskii, 2013).Atmospheric deposition of pollutants can cause detrimental effects to both terrestrial and aquatic environments and to human and wildlife health (Driscoll et al., 2001;Baldigo et al., 2009;Bobbink et al., 2010;Bacon et al., 2013;Vodyanitskii, 2013;de Vries et al., 2014;Jones et al., 2014;de Vries et al., 2015;Duan et al., 2016;Rodríguez-Estival and Smits, 2016).Once deposited onto land and water surfaces, pollutants are incorporated into soil or water where they can be taken up directly by biota, methylated by microorganisms in the case of mercury (Hg), or increase the leaching of soil base cations, metals, and plant nutrients in the case of acidifying pollutants.Moreover, pollutants can also be transported from where they were initially deposited via soil erosion, surface runoff, and groundwater flows.Physical, chemical and biological variables can affect pollutant uptake and toxicity in biota.Exposure to pollutants can lead to wide-ranging ecological consequences.Methylmercury is a neurotoxin capable of bio-accumulating in wildlife and biomagnifying in food webs; humans are typically exposed via fish or other wildlife consumption (Wiener et al., 2012).Depletion of plant nutrients and leaching of toxic inorganic aluminum due to acid rain can damage vegetation, increase stress to and reduce tolerance of plants, and inhibit tree growth.Acidified lakes are also detrimental to aquatic wildlife (Driscoll et al., 2001).Polycyclic aromatic compounds (PACs) and its subgroup of polycyclic aromatic hydrocarbons (PAHs) comprise hundreds of hydrocarbons, some of which are cancer-causing agents or capable of causing damage to DNA or reproductive impairment (Kim et al., 2013;Abdel-Shafy and Mansour, 2016).
Despite the knowledge that deposition occurs, monitoring of emissions, ambient air, and deposition is often conducted separately from the monitoring of ecosystem effects leading to gaps in the understanding of sources and fate of pollutants in the environment (Ochoa-Hueso et al., 2017).In the former case, researchers are able to track the variability in the emissions, air concentrations and deposition but are uncertain about the post-deposition effects and how these changes will impact ecosystems.In the latter, researchers are able to assess the ecosystem effects, but are uncertain what the sources of contamination are and the role of atmospheric deposition versus the physical, chemical, and biological factors influencing pollutant uptake.These knowledge gaps may be filled by establishing stronger connections between atmospheric sources or deposition and ecological effects.This can be accomplished through the identification of biomonitors and their associated effects from pollutant exposure and in the determination of the critical loads and critical levels at which these effects occur.A biomonitor is defined as a change in biological responses, ranging from molecular through cellular and physiological responses to behavioral changes, which can be related to exposure to, or toxic effects of, environmental chemicals (Peakall and Walker, 1994).A biological indicator refers to a living organism that provides information on an ecosystem.A critical load refers to a concentration limit in any media (air, water, soil, tissue).In this review paper, we examine the ecosystem and human health impacts from atmospheric deposition of acidifying pollutants, eutrophying nitrogen (N), polycyclic aromatic hydrocarbons (PAH), Hg, trace metals, and ozone (O 3 ), as well as the biological indicators that have been used to assess the health of ecosystems.Deposition studies on other groups of pollutants have also been made in literature (Lin et al., 2010;Weissengruber et al., 2018), but the information was not as rich as for those pollutants mentioned above, and thus were not included in this review.

Major Acidic Species
Major acidic species in the atmosphere include gases (SO 2 , HNO 3 , HCl, and organic acids) and particulate matters (sulphate, nitrate, chloride, organic acids), and associated aqueous-phase species (Seinfeld and Pandis, 2016).The components of acidic deposition vary based on the local and regional emissions but primarily consist of sulphur species.The relative contribution of each pollutant to total acidic deposition also depends on meteorology (e.g., amount of precipitation, sunlight, wind speed), which can impact formation rates as well as deposition velocities.Typically, oxidized sulphur (i.e., SO 2 and SO 4 2- ) and nitrogen (primarily HNO 3 and NO 3 -) dominate acidic deposition in polluted areas (Vet et al., 2014).The deposition of reduced nitrogen (NH x = NH 3 + NH 4 + ) can also acidify terrestrial ecosystems (soils) through two indirect mechanisms: (i) NH 4 + uptake into roots (which releases H + ), or (ii) nitrification to NO 3 -(which produces H + ) (Krupa, 2003).Base cation species (e.g., Ca 2+ , Mg 2+ , Na + , K + ) can mitigate acidic deposition by neutralizing acidic species, and should be taken into account when assessing acidification of receiving ecosystems (Watmough et al., 2014).The characteristics of acidifying pollutants and base cations, as well as commonly used indicators, are summarized in Table 1.

Impacts of Acidic Deposition
It is well established that acidic (sulphur and nitrogen) deposition can decrease soil pH, and increase soil solution concentrations of Al 3+ , resulting in the leaching of base cations (Ca 2+ , Mg 2+ , and K + ).Direct effects from deposition on leaves are manifested in changes in leaf physiology, including an increase in leaf roundness (Bacon et al., 2013), changes in the epicuticular and epistomatal waxes (Bartiromo et al., 2012;Elliott-Kingston et al., 2014), smaller leaves, and changes to the cuticle and stomata (Haworth et al., 2010;Haworth et al., 2012;Elliott-Kingston et al., 2014).Impacts on forests include nutrient deficiencies due to soil base cation leaching (Akselsson et al., 2007), leading to decreased growth (Savva and Berninger, 2010), changes in tree ring width (Godek et al., 2015), increased susceptibility to disease, destruction of root systems resulting in tree and crown dieback (Wilmot et al., 1995;Godek et al., 2015), and higher sensitivity to drought (Savva and Berninger, 2010).The effects of acid deposition have been found to have a geographical dependence (de Vries et al., 1995).Regional properties of an ecosystem that have been found to relate to acid deposition include soil chemistry, such as texture, pH, and humidity, and forest properties, such as tree species and age (de Vries et al., 1995;Økland et al., 2004;de Vries et al., 2007;Stevens et al., 2009a;van Dobben and de Vries, 2010).
Impacts on aquatic ecosystems from acidic deposition are fairly well understood; the mobilisation of Al from soils and decreased water pH can lead to reduced fish species richness, changes in fish health, and degraded water quality.The reduction and extinction of Atlantic salmon in the Maritimes (Watt et al., 2000) and fish populations elsewhere (Monteith et al., 2005;Gray et al., 2012) have been associated with increased water acidity.
Acidic deposition may impact wildlife by reducing the microbial biomass and shredder (macroinvertebrates that feed on plant and animal material and break it into smaller particles) biomass of gammarids (Meegan et al., 1996) and caddisflies (Simon et al., 2009), suppressing microbial respiration (Dangles and Guérold, 2001;Simon et al., 2009), depressing aquatic fungi (Clivot et al., 2014), damaging the gills of fish with gill lesions due to increases in Al and ).24.305 g mol -1 39.0983 g mol  decreases in calcium (Evans et al., 1988), and thinning of bird egg shells (Hames et al., 2002).Effects such as those mentioned above can both directly and indirectly affect macroinvertebrates (Baldigo et al., 2009;Ferreira and Guérold, 2017).Tables 2, 3, and 4 summarize the key points of these studies that examine the impacts of N, S, and soil base cation deposition, respectively.

Major N-deposition Species
Similar to acidic deposition, the major components of N-deposition vary from region to region and depend on local emissions and meteorology.Generally, N-deposition is dominated by inorganic reduced nitrogen (NH 3 and NH 4 + ) and oxidized nitrogen (NO x , HNO 3 and NO 3 -) (e.g., Harrison et al., 1999;Hayden et al., 2003;Horii et al., 2005;Zhang et al., 2009;Benedict et al., 2013;van den Elzen et al., 2018).Although organic N deposition is not typically measured, it has been shown to constitute a significant fraction (upwards of ~30%) at some field sites (Zhang et al., 2009;Benedict et al., 2013).Due to the variety of atmospheric N species, as well as technical challenges measuring certain N species (e.g., NH 3 (von Bobrutzki et al., 2010); organic N (Beem et al., 2010); HONO (VandenBoer et al., 2013)), complete accounting of total N deposition for a given ecosystem can be challenging.
Deposition of N on terrestrial and aquatic ecosystems has also resulted in direct and indirect effects on animal biodiversity due to changes in habitat structure and function (Nijssen et al., 2001;Feest et al., 2014;Jones et al., 2016;Vogels et al., 2016;Maes et al., 2017), including declines in carabids and dipterals (Nijssen et al., 2001;Vogels et al., 2016), changes in habitats for lizards (Jones et al., 2016), infestations of heather beetles (Berdowski and Zeilinga, 1983;Berdowski, 1993), and changes in butterfly species traits (Feest et al., 2014;WallisDeVries and van Swaay, 2016).Other reported impacts on fauna due to N deposition include changes in food diversity, quality, and abundance, as well as decreased reproduction (Nijssen et al., 2017).

Effects Indicators of Nitrogen Deposition
Impacts of nitrogen are widely assessed using the percent cover of individual plant species (Soons et al., 2016) and the presence of an indicator species (Wilkins and Aherne, 2016).Useful plant indicators include the graminoid:forb ratio (Stevens et al., 2006;Stevens et al., 2009b) and epiphytic lichen and terricolous lichen taxa (Stevens et al., 2012) with shifts in lichen communities occurring above a critical load of 5 kg N ha -1 yr -1 (Stevens et al., 2012).Butterflies can be effective indicators for N deposition on faunal biodiversity, with an index of butterfly sensitivity to N having been created using 25 years of data based on changes in butterfly abundance (Feest et al., 2014;WallisDeVries and van Swaay, 2016).
Tissue and litter N content and vegetation N:P ratio have been shown to be good biomonitors for shrubs and mosses and other organisms (Pitcairn et al., 1998;Hicks et al., 2000;Conti and Cecchetti, 2001;Pitcairn et al., 2001;Pitcairn et al., 2002;Pitcairn et al., 2003;Mitchell et al., 2004;Pitcairn et al., 2006;McNeil et al., 2007;Edmondson et al., 2010;Caporn et al., 2014;Harmens et al., 2015;Rowe et al., 2016;Vogels et al., 2016;Du, 2017).Ouimet et al., 2006;Ouimet, 2008 Empirical critical loads of N have been established for a range of terrestrial ecosystems, based on changes to ecosystem structure and function; a critical load of 15 kg N ha -1 yr -1 has been recommended for acidic and calcareous grassland (Bobbink and Hettelingh, 2011;Henry and Aherne, 2014), 15 kg N ha -1 yr -1 for calcareous grey dunes and 10 kg N ha -1 yr -1 for acidic grey dunes (Kooijman et al., 2016).Similarly, critical levels for atmospheric N are 1 µg NH 3 m -3 for lichens and bryophytes and 3 µg NH 3 m -3 for herbaceous species (de Vries et al., 2007;Cape et al., 2009).de Vries et al. (2007) suggested critical N concentrations in soil solution for a variety of vegetation types to protect against N leaching, with concentrations ranging from 0.2 mg N L -1 for lichens and cranberries to 6.5 mg N L -1 for deciduous forests.Similarly, a general threshold for terrestrial soil C:N ratio of < 20-25:1 has been suggested as a threshold for nitrate leaching (Bähring et al., 2017).Litter N content has also been found to be an effective indicator (White et al., 1996;Pilkington et al., 2005).Bähring et al. (2017) found shoot increment to be a more sensitive indicator than soil C:N ratio.Soil pH thresholds are suggested at 3.8 for forbs and 4.5 for acidic grasslands (Field et al., 2014).Polyphenol oxidase (PPO) activity is also another biomarker of N enrichment with little difference above an application of 60 kg ha -1 yr -1 on peach trees (Edmondson et al., 2010;Falguera et al., 2012;Caporn et al., 2014).
Indicators in aquatic ecosystems include shifts in the relative abundance of diatom communities, with a threshold for wet N deposition between 1.0 and 1.5 kg N ha -1 yr -1 (Baron, 2006;Saros et al., 2011;Nanus et al., 2012;Sheibley et al., 2014); nitrate concentrations, where a threshold of 15 mg L -1 has been determined for well water (Panno et al., 2006); and the ratio of dissolved inorganic N to total phosphorus (DIN:TP), with a suggested range between 1.5 and 3.4 (Bergström, 2010;Fenn et al., 2011;de Vries et al., 2015).

Major PAH Species
The major PAH species are generally considered to be the US EPA's 16 PAH priority pollutants, which includes naphthalene, acenaphthylene, acenaphthene, fluorene, phenanthrene, anthracene, fluoranthene, pyrene, chrysene, benz [a,h]anthracene, and benzo[ghi]perylene.These PAH species were selected as priority pollutants because they are found in the environment, analytical standards and toxicity information are readily available, and many scientific investigations have focused on these 16 PAHs (Andersson and Achten, 2015).However, recent advances in analytical methods have broadened the range of analyses to include polycyclic aromatic compounds (PACs).PACs consists of the 16 PAHs as well as alkylated PAHs, unsubstituted and alkylated dibenzothiophenes, heterocyclic aromatic compounds, and PAC transformation products, which together comprise thousands of compounds  (Manzano et al., 2017;Wnorowski and Charland, 2017).Although these additional PAC species have been the subject of far fewer studies relative to unsubstituted PAHs, their impacts have been observed in air, water, and wildlife particularly near oil sands mining regions (Kelly et al., 2009;Lundin et al., 2015;Schuster et al., 2015;Zhang et al., 2015;Jariyasopit et al., 2016).Toxicity information on these compounds is still very limited; therefore, the following sections will focus on the impacts and biomonitors of PAHs.

Major Hg Species
There are three forms of atmospheric Hg including gaseous elemental Hg (GEM), gaseous oxidized Hg (GOM), and particulate-bound Hg (PBM).GEM is the dominant form of atmospheric Hg and capable of long range transport with an atmospheric lifetime of 0.5-1 year (Driscoll et al., 2013).GOM and PBM typically have atmospheric lifetimes ranging from hours to weeks (Cole et al., 2014).Hg is bidirectional in nature, with both emission and deposition occurring between the air and the surface (Wright et al., 2016).All three forms of Hg can undergo dry deposition (Wright et al., 2016); however, only GOM and PBM can also undergo wet deposition since oxidized Hg is more water soluble than GEM.

Impacts of Mercury Deposition
Mercury deposited to the Earth can undergo methylation in the presence of sulphate-reducing bacteria (SRB).During this process, oxidized Hg is converted to methylmercury (MeHg), which is more often referred to as monomethylmercury (MMHg) (Gilmour et al., 1992;King et al., 2000).Methylmercury bioaccumulates in biota and is highly toxic.Hg deposition impacts humans indirectly through the consumption of contaminated fish, wildlife, and plants that are contaminated with Hg (Meili, 1997;Meili et al., 2003).The resulting effects are risks to the neurological, immune, and reproductive systems (Harada, 1995;Takeuchi et al., 1996;Myers et al., 1998;Schoeman et al., 2009;Bose-O'Reilly et al., 2010;Fernandes Azevedo et al., 2012;Hong et al., 2012;Rice et al., 2014).Wildlife at highest risk are large predatory fish, fish-eating mammals, and fish-eating birds.Impacts on wildlife include reduced reproduction, changes to egg incubation times, behavioural changes, and neurological problems (Myers et al., 1998;Wolfe et al., 1998;Evers et al., 2005;Scheuhammer et al., 2007;Richard Pilsner et al., 2010;Penglase et al., 2014).The consumption of fish with high levels of Hg is an indirect impact of Hg deposition through bioaccumulation of MeHg in the fish and biomagnification in the food chain.Aquatic ecosystems that are acidic, nutrient-deficient, or have higher levels of dissolved organic matter (DOM) have higher concentrations of MeHg than their counterparts (Gilmour et al., 1992;Watras and Huckabee, 1994;Driscoll et al., 2013;French et al., 2014).The effects of Hg on terrestrial and wildlife ecosystems are fairly well understood (Meili et al., 2003;Evers et al., 2005;Evers et al., 2008;Caldwell et al., 2009).Mercury concentrations in Canada are higher in the eastern part of the country than the west owing to the legacy of high emission sources to the south, the preponderance of acidic lakes and soils (Depew et al., 2013, and references therein), and the retention of Hg in organic soils in lake catchments in eastern Canada (Dennis et al., 2005).

Effects Indicators of Mercury Deposition
Ecosystem indicators that can exacerbate Hg bioaccumulation are total phosphorus (TP), DOC, acid status, and pH.Critical loads for Hg (and persistent organic pollutants; POPs) are not as well developed as they are for N and S. The concentration threshold used as an indicator of Hg sensitivity for total phosphorus is 0.03 mg L -1 , with fish having higher Hg below that threshold.DOC indicator values for Arctic tundra lakes range from 4 mg L -1 to 8 mg C L -1 (Driscoll et al., 2007;French et al., 2014;Stoken et al., 2016).Ecosystem attributes which are associated with higher concentrations of bioavailable Hg include an ANC less than 100 µeq L -1 , and a pH less than 6.0.The critical limit for Hg concentration in drinking water is 6 µg L -1 (WHO, 2011).The Canadian water quality guidelines for the protection of aquatic life are 26 ng L -1 for inorganic Hg in freshwater, 16 ng L -1 in marine surface waters and 4 ng L -1 for MeHg in freshwater (CCME, 2003).

Major Trace Metal Species
The 13 major trace metal species considered as priority pollutants by US EPA are Antimony (Sb), Arsenic (As), Beryllium (Be), Cadmium (Cd), Chromium (Cr), Copper (Cu), Lead (Pb), Mercury (Hg), Nickel (Ni), Selenium (Se), Silver (Ag), Thallium (Tl) and Zinc (Zn).These metals are part of the priority pollutants list because of their toxic effects on biota and occurrence in the environment.They also have chemical standards and analytical methods available and are produced in significant quantities (US EPA, 2017).

Impacts of Trace Metal Deposition
The impacts of trace metal deposition are well documented but especially for Pb.Much of the knowledge of the impacts of trace metals such as aluminum, vanadium, and nickel to terrestrial ecosystems has come from acidification studies as decreased pH mobilises metals causing changes that include changes to the mycorrhiza and fine root systems of plants, chlorosis, dwarfing, and reduced root and shoot growth (Rosseland et al., 1990;Efroymson et al., 1997;Ewais, 1997;Wang and Liu, 1999;Vachirapatama et al., 2011).
Trace metals can accumulate in the suspended particulates and sediments in aquatic ecosystems causing serious potential risks to the health of ecosystems (Li et al., 2013).Impacts of trace metals to aquatic ecosystems include effects on gill function by copper (Sola et al., 1995;Rajkowska and Protasowicki, 2013), iron, manganese, and zinc (Rajkowska and Protasowicki, 2013); nervous systems by zinc, manganese, and iron (Baatrup, 1991;Takeda et al., 2004); and growth and reproduction rates by lead, mercury, and cadmium (Mance, 1987;Ebrahimi and Taherianfard, 2011).Aluminum has been found to be toxic to fish and invertebrates and can interfere with the metabolic, reproductive, and breathing processes of mammals and birds, with potential links to tumours (Rosseland et al., 1990;Exley et al., 1991;Bast, 1993;Sparling and Lowe, 1996;Slaninova et al., 2014).Cadmium has been linked to changes in growth and reproduction of biota (Eisler, 1985;Gallego et al., 2012) and earthworms (Will and Suter, 1995;Chen et al., 2017).Nickel accumulation has been linked to changes in metabolism, bone densities, growth, and survival in birds (Outridge and Scheuhammer, 1993;Eisler, 1998).Histopathological changes and other effects have been observed in carp due to trace metals (Vinodhini and Narayanan, 2009;Georgieva et al., 2014).In humans, trace metal toxicities include damage to the lungs and nervous systems by Al; damage to the lungs, kidneys, and bones and renal dysfunction by Cd; and damage to the lungs, nasal passages, and skin by Ni (Williams and Burson, 1985;Denkhaus and Salnikow, 2002;Krewski et al., 2007;Trzcinka-Ochocka et al., 2010;Zambelli and Ciurli, 2013).

Impacts of Ozone Deposition
The impact of high ozone concentration on vegetation was traditionally considered as an exposure effect (concentrationbased approach) (Fuhrer et al., 1997).In recent decades, the accumulated stomatal flux (dose-based approach) was considered to be more suitable because ozone damage to vegetation is caused by injury to internal plant issue and subsequent reduction in photosynthesis, plant growth, and productivity (Musselman et al., 1994;WHO, 2000;Filella et al., 2005;Ferretti et al., 2007).An example of the comparison of the different approaches was shown in Zhang et al. (2006).This section focuses on effects of O 3 deposition on vegetation.Ozone is a highly reactive oxygen species and impacts human and animal health, as well as terrestrial ecosystems (Chappelka and Samuelson, 1998;Ashmore, 2005;Ashmore et al., 2006;Emberson and Buker, 2011;Mills et al., 2011;Fuhrer et al., 2016;Bergmann et al., 2017, and references therein).In mammals, significant impacts on respiratory health include reductions in lung function and changes in lung structure (Mckee and Rodriguez, 1993;Chen et al., 2007;Poursafa et al., 2011;Bergmann et al., 2017).Impacts to terrestrial ecosystems include decreased net photosynthesis in trees (Wittig et al., 2009), decreased tree biomass (Karlssen et al., 2003;Wittig et al., 2009), decreased grassland productivity (Volk et al., 2006), reduced flowering and bulb growth in woodland ground flora species (Keelan, 2007), plant cell death (Vainonen and Kangasjärvi, 2015), reduced cover of grass species (Thwaites et al., 2006), decreased yields of rice and winter wheat (Feng et al., 2003), agricultural yield losses (Van Dingenen et al., 2009), changes to litter (Fuhrer et al., 2016), changes in the decomposition rate of litter (Lindroth, 2010), among many others (Matyssek and Sandermann Jr., 2003).Rather than provide a table for O 3 , the reader is referred to an extensive review provided by Bergmann et al. (2017) that examines over 450 references of plant exposure to O 3 in literature.In addition, a review of impacts of O 3 on terrestrial biodiversity and their downstream effects highlights the aboveground and belowground effects of O 3 exposure (Fuhrer et al., 2016).

Effects Indicators of Ozone Deposition
Biomonitors for ozone have been studied intensively using field studies, such as the high use of white clover in ozone gardens (Villányi et al., 2008;ICP Vegetation, 2017).There are, however, sometimes difficulties in distinguishing between the impacts being due to O 3 or other pollutants, such as reactive N (Fuhrer et al., 2016).Useful biomonitors for O 3 deposition include the highly-sensitive Myrtaceae and Salicaceae families (Bergmann et al., 2017) and whereas lichens, on the other hand, are not effective biomonitors due to their high O 3 tolerance (Bertuzzi et al., 2013).An interesting ranking of native herbaceous and woody plant species can be found in Bergmann et al. (2017).Visible indicators on plants have been found to be leaf browning (Panigada et al., 2009) and necrotic lesions on leaf surfaces, such as on the upper leaf surfaces of tobacco cultivars (Ashmore et al., 1978;Ribas et al., 1998;Klumpp et al., 2006;Kafiatulla et al., 2012;Vainonen and Kangasjärvi, 2015).Other non-visible biomarkers for plants that have been found to show potential are stress transcripts, proteins, and metabolites (Sandermann, 2000).In humans, systemic responses such as inflammation and oxidative stress have been found to be the most useful (Goodman et al., 2015).Recently, an atmospheric critical level concentration for O 3 of 80 ppb was suggested for two Mediterranean trees, although this was performed under controlled conditions and has yet to be tested in the field (Fusaro et al., 2017).Interestingly, in this study, N deposition was observed to counter the effects of O 3 deposition (Fusaro et al., 2017).

CONCLUSIONS AND RECOMMENDATIONS
The ecosystem and human health impacts of the atmospheric deposition of various pollutants, and the indicators and biomonitors that have been found to be effective for these pollutants are summarized in this review paper.The impacts of the deposition of acidifying pollutants, eutrophying N, PAHs, Hg, Pb, and O 3 have been welldocumented, whereas less is known about the impacts of alkylated and heterocyclic PAC and trace metals, such as iron and manganese.Further studies on possible biomonitors for PAC and O 3 are needed.Many studies have been tested under controlled conditions only.Future work should include testing suggested effects indicators, such as the critical load of O 3 on Mediterranean trees, in the field.The current understanding of effects and indicators is due to close collaboration between various research communities.This integration is essential to addressing environmental problems.For example, it would be interesting to explore the possibility of extending the use of the STressor-Ecological Production function-final ecosystem goods and Services (STEPS) Framework, which has been applied to aquatic eutrophication (Rhodes et al., 2017), to other chemical species, such as PAHs, Hg, and trace metals.Recent success by Fernie et al. (2018) and Cruz-Martinez et al. (2015) on fecal and muscle sampling of PAHs and other contaminants in nesting tree swallows is being built upon with continuing research on toxicity in these birds.Perhaps expanding the list of chemical species, such as adding Hg, in these fecal samples could be performed concurrently to expand our knowledge of these wild birds and the effects of the Alberta Oil Sands on them and to develop this species as a potential biomonitor for future studies.geographic parameters.J. Environ.Manage.156: 52-61.Fernie, K.J., Marteinson, S.C., Chen, D., Eng, A., Harner, T., Smits, J.E.G., Soos, C. (2018).Elevated exposure, uptake and accumulation of polycyclic aromatic hydrocarbons by nestling tree swallows (Tachycineta bicolor) through multiple exposure routes in active mining-related areas of the Athabasca oil sands region.Sci.Total Environ. 624: 250-261. Ferreira, V. and Guérold, F. (2017).Leaf litter decomposition as a bioassessment tool of acidification effects in streams: Evidence from a field study and meta-analysis.Ecol.Indic.79: 382-390.Field, C.D., Dise, N.B., Payne, R.J., Britton, A.J., Emmett, B.A., Helliwell, R.C., Hughes, S., Jones, L., Lees, S. and Leake, J.R. ( 2014).The role of nitrogen deposition in widespread plant community change across seminatural habitats.Ecosystems 17: 864-877.Figueira, R., Sérgio, C. and Sousa, A. (2002).Distribution of trace metals in moss biomonitors and assessment of contamination sources in Portugal.Environ.Pollut. 118: 153-163. Filella, I., Peñuelas, J. and Ribas, A. (2005).Using plant biomonitors and flux modelling to develop O 3 doseresponse relationships in Catalonia.Environ.J.,Andrén,C.,Bishop,K.,Buffam,I.,Cory,N.,

Table 4 .
Impacts of Base Cations.