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Stratospheric Ozone and Human Health Project

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Environmental Effects of Ozone Depletion: 1994 Assessment



R. G. Zepp (USA), T. V. Callaghan (UK), and D. J. Erickson (USA)

Table of Contents

  1. Summary
  2. Introduction
  3. Terrestrial Ecosystems
  4. Aquatic Ecosystems
  5. CFC Substitutes
  6. Conclusions
  7. References


Increases in solar UV radiation could affect terrestrial and aquatic biogeochemical cycles thus altering both sources and sinks of greenhouse andchemically-important trace gases [e.g., carbon dioxide (CO2), carbon monoxide (CO), carbonyl sulfide (COS) and possibly other gases]. These potential changes would contribute to biosphere-atmosphere feedbacks that attenuate or reinforce the atmospheric buildup of these gases. Current research discussed here focuses on effects of enhanced UV-B on biological and geochemical processes in terrestrial and aquatic ecosystems.

In terrestrial ecosystems increased UV-B could modify both the production and decomposition of plant matter with concomitant changes in the uptake and release of atmospherically-important trace gases. Decomposition processes can be accelerated when UV-B photodegrades surface litter, or retarded when the dominant effect involves changes in the chemical composition of living tissues that reduce the biodegradability of buried litter. These changes in decomposition can affect microbial production of carbon dioxide and other trace gases, and also may affect the availability of nutrients essential for plant growth. Primary production can be reduced by enhanced UV-B, but the effect is variable between species and even cultivars of some crops. Likewise, the effects of enhanced UV-B on photoproduction of CO from plant matter is species dependent and occurs more efficiently from dead than living matter. The often-individualistic response of plant species to enhanced UV-B can result in changed competitive balances between co-occurring speciesplankton that produce organosulfur precursors, on bacteria that consume DMS, and on photooxidation of DMS. Early modeling efforts to simulate these interactions have appeared during the past two years.

New research on the environmental fate and impact of the hydrofluorocarbon (HFC) and hydrochlorofluorocarbon (HCFC) substitutes for CFCs has focused on trifluoroacetic acid (TFA), a tropospheric oxidation product of certain HFCs and HCFCs. These results indicate that TFA, although it may become globally distributed with increased usage of alternative fluorocarbons, is not likely to accumulate in soils and organisms. Although resistant to chemical degradation, very recent evidence indicates that TFA can be broken down by microorganisms.

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The term "biogeochemical cycles" is used here to refer to the complex interaction of biological, chemical, and physical processes that control the exchange and recycling of matter and energy at and near the Earth's surface. Research on biogeochemical cycles focuses on the transport and transformation of substances in the natural environment. Global biogeochemical cycles strongly influence atmospheric composition through their effects on the biospheric uptake and release of greenhouse gases and gases that actively participate in atmospheric chemical reactions. The latter will be referred to as "chemically-active" gases in this chapter. On the other hand, the Earth's climate and the nutrients derived from atmospheric deposition are of great importance to the sustainability of the biosphere. Atmospheric composition, climate, and the biosphere are coupled by strong interactions, including feedbacks that reinforce or attenuate climate change. A large fraction of most greenhouse and chemically active gases in the atmosphere is derived from biogeochemical processes in terrestrial and aquatic ecosystems [IPCC,1992].

Solar radiation, directly or indirectly, provides the primary driving force for biogeochemical cycles. Most of the solar radiation that reaches land or water is converted into thermal energy, but a significant part, especially that in the ultraviolet and visible region, is diverted into photochemical and photobiological processes that affect global biogeochemical cycles. Therefore, these processes are sensitive to changes in ground-level solar radiation that result from global changes in stratospheric ozone, cloud cover, aerosols, and other factors. Declines in stratospheric ozone and, by implication, increases in solar UV-B radiation reaching the Earth's surface have been particularly pronounced during the past few years. These changes, coupled with a number of recent findings that document the effects of UV-B radiation on biogeochemical cycles, have prompted this new section in the UNEP report on Environmental Effects of Ozone Depletion.

The principal goals of this section are to describe recent investigations of the effects of changing solar UV-B radiation on terrestrial and aquatic biogeochemical cycles, but we note review papers in the text that provide useful background information on this subject. Research related to biogeochemical cycles -- such as effects of UV-B radiation on photosynthesis, plant physiology, and aquatic trophic dynamics -- are reviewed elsewhere in this report. In addition, excellent overviews of global biogeochemical cycles are available as background to this section [Butcher et al., 1992; Schlesinger, 1991]. The cycles of various elements are discussed separately within this report, but it should be emphasized that biogeochemical cycles are tightly interwoven and are subject to significant feedback interactions, including those that affect ozone concentrations and ground-level UV-B radiation.

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Terrestrial Ecosystems

Carbon and Nitrogen Cycles

Plant responses

Given the central role of plant biology in biogeochemical cycling, understanding UV-B effects on plants is of critical importance. Here we briefly consider aspects of these effects that are relevant to biogeochemical cycles; a much more detailed discussion of plant responses appears in Chapter 3. Most studies of the interactions of UV-B and plants have been conducted under artificial light conditions in growth chambers or greenhouses [Tevini and Teramura, 1991; Krupa and Kickert, 1989; SCOPE,1992;1993] that may not mimic the spectral characteristics of ozone-dependent UV-B changes occurring in solar radiation that reaches the ground. However, these studies have provided insights into general UV-B effects on physiological responses of higher plants and on alteration of growth rates and yields of crop plants. Fewer studies address growth responses of long-lived species or primary production and phytomass in natural ecosystems. As discussed in more detail in Chapter 3 of this report, responses to UV-B radiation are variable between species and even between cultivars of the same species. In general, however, it seems that those species experiencing naturally high fluxes of UV-B radiation (e.g., alpine and tropical) are better adapted to withstand increases in UV-B than those from regions currently exposed to low levels of UV-B. For example, growth chamber studies of various plants from the Arctic and lower latitude alpine regions indicated that the arctic species, living where UV-B radiation is lower, exhibited significantly higher photosynthetic inhibition than the low-latitude species where solar UV-B flux is high [Caldwell, et al., 1982].

The few field studies that have been conducted indicate lower plant sensitivity to UV-B than that observed under growth chamber and greenhouse conditions [SCOPE, 1993], probably in part because of the higher flux of photorepairing solar radiation in the UV-A and other spectral regions and the plant development effects that are prevalent in the field. Multi-season field studies of loblolly pines, however, have shown statistically significant reductions of growth under enhanced UV-B radiation that corresponded to 16% and 25% decreases in total ozone [Sullivan and Teramura, 1991]. Moreover, the UV-B effects on growth accumulated over 3 years of exposure. Photosynthesis quantum yields were generally reduced in the loblolly experiments and the effect was attributable to direct effects on photosystem II. Field studies on Swedish subarctic heath vegetation indicated greater reductions in growth of shoots in two dominant evergreen dwarf shrubs which accumulated damage, than in two deciduous dwarf shrubs [Johanson et al., 1994].

Reviews of UV-B radiation effects on plants [Tevini and Teramura,1991; Krupa and Kickert, 1989; SCOPE, 1993] refer to publications that show these effects to be on plant morphology and flowering as well as growth and photosynthesis. Particular effects vary among species, among cultivars of a species and among local populations of a species. These changes can alter the competitive balance in mixed species stands, resulting in significant changes in species composition and thus in primary productivity. Rapid global warming is predicted to cause similar effects. Changes in primary production affect the flow of CO2 through the biosphere, but do not necessarily affect carbon storage. However, any changes in the composition of species in plant communities driven by increased UV-B -- particularly in the representation of different life forms -- could significantly affect the amount of carbon stored in phytomass. For example, any shift from UV-B sensitive evergreens to deciduous dwarf shrubs [implied from work by Johanson et al. 1994] or trees could reduce carbon storage during winter while decreasing growth in general and would increase the amount of CO2 circulating in the atmosphere. Any sensitivity of the ground layer to UV-B, particularly mosses [Gehrke, 1992; Sonesson and Callaghan, 1994], would alter their carbon storage and could also increase the temperature of soil, which they insulate, thereby stimulating microbial activity and CO2 release to the atmosphere. Stresses caused by increased UV-B radiation, in combination with climatic change, may affect species composition and herbivory and enhance susceptibility of plants to insects, disease and fire. Because high latitudes are experiencing particularly large increases in solar UV-B radiation, UV-B stresses of high-latitude forests may reinforce their transient release of carbon to the atmosphere in response to climate change [Smith and Shugart, 1993].

Likely UV-B increases due to stratospheric ozone depletion are not the only environmental changes. Increases in concentrations of atmospheric CO2 are well documented and, as they generally stimulate plant productivity, at least in the short term, the balance between UV-B and CO2 interactions could possibly be a particularly important determinant of carbon cycling.

Very little is known about the effects of enhanced solar UV radiation on the terrestrial nitrogen cycle. This lack of information is of concern because changes in terrestrial nitrogen cycling would likely affect the release of N2O, a gas that is predominantly derived from terrestrial systems [IPCC, 1992]. N2O is an important greenhouse gas and it also participates in stratospheric processes that control the ozone layer. The changes in species composition that are discussed above and in Chapter 3 would likely affect nitrogen cycling in terrestrial ecosystems, e.g., possibly via UV-B effects on symbiotic association between higher plants with mycorrhizae and nitrogen fixing bacteria in root nodules. In the above-cited study of the subarctic dwarf shrub, Vaccinium uliginosum [Gehrke et al., 1994], it was found that plants exposed to enhanced UV-B had much higher levels of leachable ammonia in the litter than the controls, indicating possible effects on ammoniafication.

Litter decomposition

For many systems, site fertility is largely dependent on decomposition processes that release nutrients bound within non-living organic matter (NLOM). Therefore, changes in plant litter production and/or subsequent degradation of litter, resulting from higher UV-B levels, may have significant impacts on nutrient cycling. Decomposition of NLOM is carried out by saprophytic fungi and bacteria.

UV-B can potentially affect litter decomposition in several ways: (1) by changing the quantity of litter that is available for decomposition; (2) by altering root/shoot ratios that determine where, i.e., below ground or on the surface, and how efficiently plant matter is decomposed ; (3) by photoinhibition of biota that decompose surface litter; (4) by photodecomposition of surface litter and; (5) by changes in chemical composition of litter that alter its microbial decomposition. In comparison to studies of effects of enhanced solar UV-B radiation on plant morphology and flowering, growth and photosynthesis, little is known about the effects on decomposition.

Recent field studies in northern Sweden confirmed that enhanced UV-B radiation had significant effects on the quality and decomposition of litter from the subarctic dwarf shrub, Vaccinium uliginosum [Gehrke et al., 1994; Jones et al., 1994]. The leaves of the shrub were richer in soluble carbohydrates and tannins than in the controls and the initial decomposition rate of the leaves mixed with soil from the field site [expressed as micrograms CO2 per gram dry weight (g DW) per hour] was reduced (Figure 5.1). Additional studies in microcosms further showed that leaf litter obtained from the shrubs grown under enhanced UV-B contained more leachable ammonia and less cellulose and lignin than did the litter from the control shrubs. Fungal communities were affected and fungal and microbial activity was also significantly reduced on the plant leaves that were exposed to enhanced UV-B, such that less CO2 was emitted to the atmosphere, i.e., soil carbon storage was increased. The changes in chemical composition of leaves may involve photoreactions of lignocellulose materials. UV irradiation of lignocellulose materials results in dimerization, oligomerization, and quinone formation through photoreactions that involve carbonyl and phenolic functional groups [see Heitner and Scaiano, 1993 for review].

In contrast to the inhibiting effects of UV-B radiation on biotic decomposition of litter, other studies indicate that direct exposure of litter to enhanced UV-B increases its decomposition rate [see Moorhead and Callaghan, 1994 for review]. Moorhead and Callaghan [1994] have used the CENTURY model to simulate potential effects of UV-B induced litter degradation on nutrient dynamics and soil carbon storage. Results of the simulations indicated that the increased surface litter degradation rate substantially decreased the surface litter and lignin pool sizes. The simulations indicated, however, that these changes had little effect on pool sizes of the passive and slow-cycling organic matter in the soil. It was concluded, therefore, that enhanced UV-B may have little effect on long-term nutrient cycles since it is through formation of resistant soil organic matter (SOM) complexes that nutrients are sequestered.

In summary, decomposition processes can be accelerated when UV-B photodegrades surface litter, or retarded when the dominant effect involves changes in the chemical composition of living tissues that reduce the biodegradability of buried litter. These changes in decomposition can affect microbial production of carbon dioxide and other trace gases, and also may affect the availability of nutrients essential for plant growth. Factors that determine the net effect are poorly understood. In those ecosystems where decomposition is retarded by increased UV-B, any additive or synergistic interactions with increasing levels of CO2 which have also been shown to retard decomposition [Couteaux et al. 1991] could lead to increased carbon storage in soils but lower primary production. Such effects are often overlooked when feedbacks from warming soils due to climate change are calculated.


Fig. 5.1. Respiration rates of microorganisms feeding on plant litter from Vaccinium uliginosum leaves grown under enhanced vs. ambient UV-B radiation in the field [Gehrke et al., 1994]. Data represent 12 replicates of experiments. Results indicate that enhanced UV-B exposure reduced the litter decomposition rate.

Trace Gas Exchange Dynamics

In the above discussions, effects of enhanced UV-B on the exchange of carbon dioxide between terrestrial ecosystems and the atmosphere were emphasized. In addition to uptake of CO2, plants are known to release the chemically important gases CO and non-methane hydrocarbons (NMHCs), to the atmosphere and to take up nitrogen oxides (NOx) and COS [IPCC, 1992].

Isoprene and other NMHCs react in the troposphere to produce ozone, other oxidants and aerosols. In addition, CO and NMHCs are important scavengers of OH radicals in the troposphere (Chapter 6) and changes in OH concentrations may affect the concentration of the greenhouse gases, methane and CFCs. The direct effects of UV-B on release of NMHCs, such as isoprene, are unknown, but it is known that biogenic emissions from vegetation are species dependent. Thus, even if direct effects of UV-B radiation have little effect, changes in the species composition of plant communities, driven by increased UV-B and climatic stressors, could affect net fluxes of these chemically-important gases.

Laboratory studies have shown that senescent and dead leaves from temperate deciduous plants and tropical grasses photoproduce CO much more rapidly than living plant leaves [Tarr et al.,1994]. Wavelength studies (>300 nm) indicate that UV-B radiation produces CO from the leaves with the highest efficiency, although UV-A radiation also induces CO formation. The action spectra vary significantly from one species to another. The flux of CO from living and non-living plant matter is sufficiently great that it may be a major global source of this chemically active gas [IPCC, 1992]. As discussed below, UV-B radiation also affects the production of CO in aquatic ecosystems.

Current research indicates that uptake by terrestrial plants is the major global sink for COS [Bates et al., 1992]. Moreover, within-canopy uptake of the NOx from soils occurs in tropical humid forests. Based on the known effects of UV-B on plant growth and photosynthesis that are discussed above, it is likely that changes in UV-B radiation would selectively affect plant uptake of COS and NOx. To date, however, there are no experimental data that confirm this possibility. COS, a gas that is considered to be one of the major sources of sulfate aerosols in the stratosphere during periods of quiescent volcanic activity [Crutzen, 1976], is discussed in more detail below. NOx, like NMHCs and CO, also participates in tropospheric chemical processes that affect concentrations of tropospheric ozone, methane, and CFCs (see Chapter 6). Thus, both COS and NOx, like CO, may be involved in feedbacks that affect ground-level UV-B radiation and climate change.

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Aquatic Ecosystems

Carbon and Nitrogen Cycles

Responses of photosynthetic organisms

Upper ocean carbon cycling is intimately tied to both the UV-B induced stresses to extant phytoplankton communities as well as the complicated interaction of DOM with UV-B radiation. The most direct effect of increased UV-B fluxes on upper ocean carbon cycling would be through changes in phytoplankton community fecundity and structure. Other indirect effects involving trophic level interactions may also affect ecosystem productivity [Bothwell et al., 1994]. In this section we discuss the possible role of these changes in aquatic carbon cycling. Phytoplankton responses are considered in much greater detail in Chapter 4.

Laboratory and field studies with photosynthetic organisms obtained from aquatic ecosystems in different locations indicate that reductions in current levels of solar UV-B result in enhanced primary production [for recent review see Karentz et al., 1994]. Antarctic experiments under the ozone hole demonstrated that primary production is inhibited by enhanced UV-B. Figure 5.2 shows the observed decreased phytoplankton production associated with a 33% decrease in column ozone abundance during austral spring of 1990 in Antarctica [Smith et al., 1992]. The net impact of a reduction in primary production on the ocean sink for atmospheric CO2 is uncertain. Most primary production is recycled in the upper layer of the sea. Only a fraction of the upper ocean particulate organic carbon (POC) and dissolved organic carbon (DOC) actually is exported from the upper ocean into intermediate and deep water. On a global basis this "new production" is estimated to be about 10 Gt C [1 gigaton (Gt) of carbon equals 1015 g] and it is believed that most of this POC and DOC is remineralized in the top km of the sea [Siegenthaler and Sarmiento, 1993]. About 0.2 Gt C of POC reaches the bottom sediments annually where long-term carbon storage takes place. Taking into account biological removal as well as vertical water transport of the dissolved inorganic carbon (DIC), the amount of anthropogenic CO2 taken up by the ocean annually has been estimated to be about 2 Gt in recent years [Siegenthaler and Sarmiento, 1993]. Additional field and modeling studies are required to come up with reliable estimates of the impacts of enhanced UV-B radiation on this oceanic sink.

As in the case of terrestrial plants, aquatic photosynthetic organisms (algae, cyanobacteria) differ substantially in their tolerances to UV-B exposure and changes in the ratio of UV-B:UV-A:PAR [UNEP, 1991; SCOPE, 1993; Prezelin et al., 1994; Weiler and Penhale, 1994]. Evidence has been presented that the balance between photosynthesis, photoinhibition, and photorepair processes for natural communities of photosynthetic organisms has been perturbed by ozone depletion over Antarctica. This perturbation may result in changes in the competitive balance of species over time [Prezelin et al., 1994; Weiler and Penhale, 1994]. A recent study has indicated, however, that compositional changes in the diatom component of the Antarctic phytoplankton community over the past 20 years have cannot be differentiated from natural variability [McMinn et al., 1994].

Marine phytoplankton produce NMHCs [Bonsang et al., 1988; Donahue and Prinn, 1990], but the effects of UV-B radiation on their photoproduction is unknown. As noted earlier in this chapter and elsewhere in this report, these compounds interact with the hydroxyl radical [Donahue and Prinn, 1990] which in turn plays a dominant role in maintenance of the oxidizing capacity of the atmosphere. Phytoplankton, as well as higher plants and bacteria, produce other non-volatile hydrocarbons that are highly resistant to biodegradation [for review see de Leeuw and Largeau, 1993], but the effects of enhanced UV on production of these biomacromolecules have not been examined.


Fig. 5.2. Average values for in situ phytoplankton productivity versus depth (m) within the marginal ice zone of the Bellinghausen Sea in austral spring of 1990 in which reduced productivity inside ozone hole is compared with productivity outside hole. Integrated over depth in the water column, reductions in productivity ranged from 6 to 12%. On an annual basis, this range corresponds to an estimated annual productivity loss of 7 to 14 teragrams which is 2 to 4% of production in the Antarctic marginal ice zone and about 0.1% of global phytoplankton production [Smith et al., 1992].

A large fraction of the sea surface is covered by an organic microlayer that may have significant effects on air-sea gas exchange. The sea surface microlayer is fully exposed to solar UV radiation and thus may be particularly susceptible to effects of enhanced UV radiation [Duce and Liss, 1995].

As in the case of terrestrial ecosystems, few data are available on the effects of UV radiation on aquatic biota involved in the nitrogen cycle. Changes in species composition noted above are likely to affect nitrogen cycling, and suppression of assimilatory nitrate reduction by phytoplankton has been demonstrated [UNEP, 1991]. The effects of UV-B on decomposition processes that are discussed below also are likely to perturb aquatic nitrogen cycling. Moreover, solar UV-B radiation is mainly responsible for the photodegradation of nitrate in water [Zafiriou and True, 1979; Zepp et al., 1987]; among other species, hydroxyl radicals are produced in this photoreaction. Nitrite is also degraded by solar UV radiation to form nitric oxide (NO), hydroxyl radicals, and other products, but this reaction is mainly induced by UV-A radiation [Zafiriou and Bonneau, 1987]. Effects of UV-B on marine nitrogen cycling could affect sea-to-air exchange of N2O and NO. However, the role of the ocean as a source of atmospheric N2O is poorly defined [IPCC, 1992; Butcher et al., 1992].


Enhanced UV radiation may affect the decomposition of POC and DOC through its effects on bacterial activity and through photodegradation of DOC. Freshwater and marine bacteria from both freshwaters and the sea are impacted by changes in solar UV radiation [for review see Karentz et al. ,1994] and action spectra indicate that UV-B radiation is mainly involved [Calkins and Barcelo, 1982]. This has been confirmed by field studies which indicate that current levels of solar UV radiation reduce bacterioplankton growth in the upper ocean [Herndl et al., 1993]. This effect may retard the decomposition of labile organic matter in the upper ocean. Effects on other biogeochemical cycles are discussed below.

That aquatic DOM is photoreactive has been established by numerous studies [Kouassi et al., 1990, 1992; Frimmel and Bauer, 1987; Kieber et al., 1990; Francko, 1990; Mopper et al., 1991; Valentine and Zepp, 1993; Miller and Zepp, 1994]. Evidence for such photoreactivity derives in part from observed changes in the electronic absorption spectrum of the organic matter as well as in the fluorescence intensity and spectrum. The sunlight-induced decrease in absorbance is not accompanied by a corresponding change in dissolved organic carbon content (DOC), although conversion to various UV-transparent products does occur (see below). These spectral changes involve greater bleaching of the ultraviolet part of DOM spectra than the visible part [Kouassi et al., 1992; Miller and Zepp, 1994]. Thus, photodegradation of DOM may result in deeper penetration of solar UV (compared to visible) radiation into the sea and freshwaters.

The photoinduced fading of DOM is accompanied by the formation of a varietyof organic and inorganic compounds. The photoproducts of the biologically refractory DOM in natural waters are DIC [Miller and Zepp, 1994]; low molecular weight compounds that are biologically labile, e.g., formaldehyde, acetaldehyde, and the alpha-keto acid, glyoxylate [Kieber et al., 1989]; the trace gases, CO and COS; and other unidentified species. Photoproduction rates of these compounds are greatest near the surface of inland and coastal waters and least for open ocean waters. Francko [1990] has reviewed the rather sparse literature on effects of solar radiation on the bioavailability of DOM. DIC is the major product from the photodegradation of DOM derived from a coastal estuary and two rivers, including DOM in the Mississippi River plume, Gulf of Mexico [Miller and Zepp, 1994]. The formation rate of DIC is at least an order of magnitude greater than that of other photoproducts. Mopper et al. [1991] have argued that photodegradation may limit the lifetime of biologically refractory DOM in the ocean.

UV-B radiation was mainly responsible for the photodegradation of DOM in Biscayne Bay (Miami), Florida, other coastal waters and the open sea [Kieber et al., 1990; Mopper et al., 1991]. and also induces the photoproduction of COS in near coastal seawater [Zepp and Andreae, 1994] (Figure 5.3). Although solar UV radiation is predominantly responsible for DOM photodegradation, the action spectra vary from one region to another and action tails well into the UV-A region in some cases [Valentine and Zepp, 1993; Zepp and Andreae, 1994] (Figure 5.3).

Photochemical reactions of DOM in the surface ocean produce dissolved gases that are supersaturated with respect to an equilibrium state with observed atmospheric concentrations. This imbalance, or 'dis-equilbria', drives the flux of these trace gases to the atmosphere from the ocean. In order to predict quantitatively the effects of decreases in stratospheric ozone and enhanced UV-B radiation on the photoproduction of these gases, action spectra are required. The air-sea exchange of gases also is affected by changes in wind speed caused by climate change. Wind speed affects sea-to-air transfer coefficients as well as vertical mixing in the upper ocean. To provide a framework for estimation of the flux and distribution of CO and COS in the upper ocean, Najjar et al. [1994] have described a model that takes into account photoproduction, turbulent mixing and chemical and biotic sinks for CO and COS in the upper ocean.

DOM photoreactions are believed to be the main source of CO in seawater; its loss has been ascribed primarily to microbial metabolism. As a result of these two processes, CO emissions from the sea follow a diurnal pattern with maximum near surface ocean concentrations greatly exceeding saturation during daylight. Although the sea is thought to be a net source of CO, great uncertainty exists regarding the strength of this source. Estimates range from 10 Tg/y up to 200 Tg/y [Erickson, 1989; IPCC, 1992]. Photodegradation of DOM in wetlands and near coastal waters may be an important regional source of CO [Valentine and Zepp, 1993]. Action spectra and quantum yields for CO photoproduction were found to be similar for water obtained from several wetland and near coastal sites in North America. For wavelengths > 300 nm, the greatest action for CO production was in the UV-B region, but the spectra tailed out well into the UV-A region.

Interrelationships among biogeochemical processes in the ocean and atmosphere can sometimes lead to feedbacks. Such a feedback relationship involving marine CO and tropospheric ozone illustrates a possible feedback between ozone depletion and air-sea exchange of trace gases [Erickson, 1989].


Fig. 5.3. Comparison of action spectra for the photoproduction of COS inseawater obtained from coastal regions of the North Sea (m) and Gulf of Mexico (n). North Sea water was obtained near Bremerhaven, Germany ( 54°N, 8°E) and Gulf of Mexico water near Turkey Point, Florida (30°N, 84.5°W). Samples used in the action spectra studies were studied within 12 hours of the time of collection [Zepp and Andreae, 1994].

Sulfur Cycle

Increased UV-B can influence the sulfur cycle (Figure 5.4) via impacts on both aquatic and terrestrial ecosystems. Changes in the sea-to-air transfer of DMS and COS may influence the radiative balance of the atmosphere. As noted in a previous section, the growth and productivity of oceanic phytoplankton as well bacterioplankton growth may be affected by a variety of stressors including increased UV-B. Production of dimethylsulfonium propionate (DMSP), the precursor compound of DMS, by phytoplankton such as coccolithophorids provides the primary source (up to 90%) of sulfur for cloud condensation nuclei in the remote marine atmosphere. On a global scale, marine emissions of DMS account for about 15% of the total atmospheric sulfur input [Bates et al., 1992]. Because the main sources of DMS in seawater are particular species of phytoplankton, any alteration in the fecundity and species distribution of phytoplankton communities could have a direct effect on the surface ocean DMS concentration and subsequently the sea-to-air flux of DMS. In addition to its effects on phytoplankton, UV-B radiation may affect the loss of DMS through microbial [Kiene and Bates, 1990] and photooxidative degradation. UV-B induced changes in these various upper oceanic processes that affect DMS may result in changes in net sea-to-air flux.

Field studies have indicated that a variety of processes affect the atmospheric concentrations of COS [Andreae and Ferek, 1992]. These studies show that COS is formed primarily by photochemical processes in the upper layers of the ocean. Zepp and Andreae [1994] have found that the photochemical formation of COS from dissolved organosulfur compounds in sea water can be photosensitized by DOM. Because rates of photosensitized reactions are generally much more rapid in coastal waters than in the open sea [Kieber et al., 1990; Zafiriou and Dister, 1991], these results help to explain why concentrations of COS have been observed to be highest in coastal regions [Andreae and Ferek, 1992]. Global estimates of COS production recently have been derived using Coastal Zone Color Scanner satellite data and general circulation models [Erickson and Eaton, 1993]. Wavelength studies of COS formation in coastal seawater samples have confirmed that COS is predominantly formed by the action of middle UV radiation (280-340 nm), but that action spectra for the North Sea and Gulf of Mexico differed significantly beyond 320 nm [Zepp and Andreae, 1994] (Figure 5.3).


Fig. 5.4. Biogeochemical processes affecting the global sulfur cycle [adapted from SCOPE, 1993]

Oxygen Cycle

The previously-discussed effects of UV-B radiation on photosynthesis and microbial decomposition also can affect the oxygen cycle in both aquatic and terrestrial ecosystems. Photooxidation of marine and freshwater DOM leads to consumption of oxygen as it is combined with the DOM carbon. DOM photooxidation also is accompanied by reduction of oxygen to form superoxide [Zafiriou and Dister, 1991], which dismutes (i.e., disproportionates) to form hydrogen peroxide. Other reactive oxygen species are produced on absorption of UV radiation by freshwater and marine chromophores [see Waite et al., 1988 and Blough and Zepp, 1994, for reviews]. Hydrogen peroxide, through interactions with marine biota and chemical constituents, is oxidized back to oxygen or reduced to water. Action spectra for the photoproduction of hydrogen peroxide in sea water and freshwaters indicate that solar UV radiation is most effective [Moore et al., 1993]. Action is most pronounced in the UV-B region, although UV-A radiation also is involved. A model that describes the upper ocean distribution of hydrogen peroxide has been developed [Sikorski and Zika, 1993].

Metals Cycles

Field and laboratory studies over the last 5 years have indicated that solar UV radiation enhances the reductive dissolution of iron and manganese oxides/hydroxides in oxygenated natural waters [Faust, 1994; Sulzberger, 1994; Waite et al., 1994]. Reductive dissolution converts the thermodynamically stable, but biologically unavailable, oxides of these trace metals into more bioavailable forms, and thus may help control marine productivity in parts of the ocean that are limited by iron or manganese. The action spectra for iron photodissolution processes have not been determined and likely are variable. Light enhances the dissolution of an amorphous manganese oxide in seawater and this process helps account for the surface maxima in dissolved manganese (Mn[II]) observed in the ocean [Sunda and Huntsman, 1990]. UV-B radiation inhibits the microbial oxidation of soluble Mn[II] to low-solubility Mn[IV] oxides [Sunda and Huntsman, 1990]. In most of these cases, the presence of aquatic DOM was shown to be essential for the occurrence of photodissolution. That naturally occurring organic compounds are capable of inducing or assisting the photodissolution of iron and manganese oxides has been confirmed in a number of laboratory studies using organic acids that are present in fresh and marine waters, e.g. humic, fulvic and hydroxycarboxylic acids [Waite et al., 1994].

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CFC Substitutes

Considerable research on the development of suitable CFC substitutes has taken place during the past few years. This research has included intense studies of the atmospheric fate and impact of HFC and HCFC substitutes for CFCs [see AFEAS, 1994 and Wallington et al., 1994 for reviews]. In addition to evaluations of the ozone depletion potentials of these substitutes, studies have included initial evaluations of potential impacts of deposition of selected atmospheric degradation products of HCFCs and HFCs. TFA, the major atmospheric degradation product of HCFC-123, HCFC-124, and HCFC-134a was selected for study, because it is resistant to abiotic degradation, including atmospheric decomposition, and likely to become globally distributed. Measurements of TFA in the present environment are sparse. During late 1993 and early 1994, TFA was detected in the atmosphere near Tuebingen, Germany and on spruce needles in Sweden. It is most likely that the source of this TFA was atmospheric degradation of the anesthetic, halothane. Models indicate that most of the TFA formed from HCFCs and HFCs should be deposited in the ocean, where it will be diluted to very low concentrations. Up to 30% may be deposited on land. Recent field studies have shown that TFA, a hydrophilic substance, is not likely to accumulate in most soils or biota, although accumulation in acidic organic soils may be possible. Evidence is now emerging that TFA may be degradable by microorganisms [Visscher et al., 1994] and that, even at high concentrations, it appears not to significantly inhibit the metabolism of acetate. Additional research is currently underway to better elucidate the biospheric fate and impacts of TFA and potential effects of TFA on biogeochemical cycles.

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Potential effects of enhanced UV-B radiation on terrestrial and aquatic carbon, nitrogen, sulfur, oxygen, and metal cycles have been identified. These effects and companion effects on biogeochemical cycles in the atmosphere may result in feedbacks that either reinforce or attenuate the buildup of greenhouse gases and aerosols in the atmosphere. Radiation amplification factors shown in Chapter 1 for some of the aquatic photochemical processes discussed in this section indicate that they are approximately as sensitive to changes in the stratospheric ozone layer as the effects on health, plants, and tropospheric photolysis.

Increasing UV-B radiation has the potential to change the quantities and chemical species of carbon that are exchanged between the atmosphere and terrestrial biosphere and also the carbon that is stored in soils and aquatic systems. The effects of UV-B on carbon storage and carbon fluxes on land are complex, varying between ecosystems, species and even crop cultivars. In some systems, UV-B can, for example, increase carbon storage in soils, whereas in others it can enhance the degradation of lignin in soil organic matter and the photoproduction of chemically-important gases such as carbon monoxide.. Responses of plants to UV-B have been shown to increase over time; impacts on natural ecosystems may be subtle but may lead to species shifts in the long term. Consequently, there is a need for research on the impacts of UV-B radiation on biogeochemical cycling to develop a new focus, in addition to investigations to determine short term physiological responses. This focus should address long-term productivity responses under natural conditions.

Laboratory and field experiments in aquatic ecosystems have shown that enhanced UV-B has a variety of effects on biogeochemical cycles. For example, photosynthesis has been suppressed by ozone depletion and enhanced UV-B radiation over the Antarctic. Moreover, recent studies indicate that photodegradation of DOM in the upper ocean has multiple effects on biogeochemical cycles, ranging from changes in the penetration of UV-B into the sea to enhanced formation of DIC and CO. Sulfur cycling in marine systems also might be affected by changes in UV-B radiation with resulting effects on the sea to air fluxes of COS and DMS. Metal cycles and bioavailability, especially those of the trace nutrients, iron and manganese, are sensitive to UV changes as well. Models designed to simulate these effects and their interactions with changing climatic conditions are just beginning to be developed.

The results discussed here suggest that changing UV-B radiation may affect global biogeochemical cycles and related feedback interactions. It should be emphasized that evaluations of regional and global scale effects require the development of appropriate models and observational approaches. Earth system models that incorporate solar UV-B radiation as a forcing variable are required in order to integrate, evaluate and predict ecosystem effects and related feedbacks on a global scale.

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