-------
PQ
VI
1
8
o
VI
s
v
*-H
O
o
#
s
13 -9 -ii 'a "§ -S
•a fl & S ^ o
1 § s I ^ a
ilgg
August 1991
9B-32 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
10. THE EFFECTS OF NITROGEN OXIDES ON
NATURAL ECOSYSTEMS AND THEIR COMPONENTS
10.1 INTRODUCTION
The previous chapter discusses the responses of individual plants exposed to nitrogen
oxides (NO, NO^. This chapter explains the known effects of nitrogen compounds (e.g.,
nitrogen oxides, nitrate, nitric acid) on terrestrial and aquatic communities. Because the
various ecosystem components are chemically interrelated, stresses placed on the individual
components, such as those caused by nitrogen loading, can produce perturbations that are not
readily reversed and will significantly alter an ecosystem.
Since the mid-1980s the view has emerged that the atmospheric deposition of inorganic
nitrogen has impacted aquatic and terrestrial ecosystems. It is known that in many areas of
the United States the atmospheric input of nitrogen compounds is significant (U.S.
Environmental Protection Agency, 1982), however, the impacts are generally unknown or
considered benign. Although, the evidence linking nitrogen deposition with ecological
impacts is tenuous there has been a growing concern (Skeffmgt6n and Wilson, 1988). This
concern has,been magnified because (1) the atmospheric concentrations of nitrogen
compounds have increased in North America and most European countries, and
(2) ecosystems formerly limited by nitrogen have .become nitrogen saturated via atmospheric
deposition. These concerns have led to attempts to develop "critical loads" of nitrogen for
various ecosystems. A "critical load" is defined as, "a quantitative estimate of an exposure to
one or more pollutants below which significant harmful effects on specified sensitive elements
of the environment do not occur according to present knowledge" (Nilsson and Grennfelt,
1988).
This chapter is organized into six main sections that are presented in the following
sequence: (1) overview and description and responses of ecosystems to impairment of
functions; (2) a generalized description of the nitrogen cycle; (3) deposition of nitrogen into
ecosystems; (4) terrestrial ecosystem effects, specifically the response ,of soil and vegetation
to nitrogen deposition; (5) effects on wetlands and bogs of nitrogen loading; and
(6) discussion of the effects on aquatic ecosystems of nitrogen loading.
August 1991
10-1
DRAFT-DO NOT QUOTE OR CITE
-------
1 10.1.2 Ecosystems
2 Ecosystems are composed of populations of "self-supporting" and "self maintaining"
3 living plants, animals and microorganisms interacting among themselves and with the non-
4 living chemical and physical environment within which they exist (Odum, 1989; Billings,
5 1978; Smith, 1980). Ecosystems usually have definable limits and may be large or small
6 (©•§•> fallen logs, forests, grasslands, cultivated or uncultivated fields, ponds, lakes, rivers,
7 estuaries, oceans, the earth) (Odum, 1971; Smith, 1980; Barbour et al., 1980). The
8 environmental conditions of a particular area or region determine the boundaries of the
9 ecosystem as well as the organisms that can live there (Smith, 1980). Together, the
10 environment, the organisms and the physiological processes resulting from their interactions
11 form the life-support systems that are essential to the existence of any species on earth,
12 including man (Odum, 1989).
13 Human welfare is dependent on ecological systems and processes. Natural ecosystems
14 are traditionally spoken of in terms of their structure and functions. Ecosystem structure
15 includes the species (richness and abundance), their mass and arrangement in an ecosystem.
16 This is termed an ecosystem's standing stock—nature's free "goods" (Westman, 1977).
17 Society reaps two kinds of benefits from the structural aspects of ah ecosystem: (1) products
18 with market value such as fish, minerals, forest products and Pharmaceuticals, and genetic
19 resources of valuable species (e.g., plants for crops and timber, and animals for
20 domestication), and (2) the use and appreciation of ecosystems for recreation, esthetic
21 enjoyment, and study (Westman, 1977).
22 More difficult to comprehend, but no less vital, are the functional aspects of an
23 ecosystem. They are the dynamics of ecosystems and impart to society a variety of benefits,
24 nature's free "services." Ecosystem functions encompass the interactions of its components
25 and their environment and maintain clean air, pure water, a green earth arid a balance of
26 creatures; the functions that enable humans to obtain the food, fiber, energy and other
27 material needs for survival (Westman, 1977). •• •••:•
28
29 10.1.2.1 Characteristics of Ecosystems
30 Ecosystems have both structure arid function. Structure within ecosystems involves
31 several levels of organization. The most visible are: (1) the individual and its environment;
August 1991
10-2
DRAFT-DO NOT QUOTE OR CITE
-------
1 (2) the population and its environment; and (3) the biological community and its
2 environment, the ecosystem (Billings, 1978). The responses of the constituent organisms to
3 environmental changes or perturbations determines the response of the ecosystem.
4 Populations of plants, animals, and microorganisms (producers, consumers, and decomposers)
5 "within, an ecosystem live together and interact as communities. Communities, due to the
6 interaction of their populations and of the individuals that constitute them, respond to
7 pollutant stresses differently from individuals. Organisms vary in their ability to withstand
8 environmental changes.^ The range of variation within which individual organisms can exist
9 and function determines the ability of a population of organisms to survive.
10 Intense competition among plants for light, water, nutrients, and space, along with
11 recurrent natural climatic (temperature) and biological (herbivory, disease) stresses can alter
12 the species composition of communities by eliminating those individuals sensitive to specific
13 stresses. Those organisms able to cope with the stresses survive and reproduce. Competition
14 among plants of the same species does not influence species succession (community change
15 over time). Competition among different species, .however, results in succession and
16 ultimately produces ecosystems composed of plant species that have a capacity to tolerate the
17 competitional stresses (Kozlowlski, 1980). Pollutant stresses are superimposed upon the
18 naturally occurring competitional stresses mentioned above. Air pollutants are known to alter
19 the diversity and structure.of plant communities (Guderian et-al., 1985). The primary effect
20 of air pollutants is on the more susceptible members of the plant community in that they can
21 no longer compete effectively for essential nutrients, water, light, space etc. As a
22 consequence of altered competitive conditions in the community, there is a decline in the
23 sensitive species permitting the enhanced growth of more tolerant species. The extent of
24 change that may occur in a community depends on the condition and type of community as
25 well as the pollutant exposure. s '
26 .-....;..
27 10.1.2.2 Ecosystem Functions
28 Ecosystem function refers to the suite of processes and interactions among its
29 components and their environment that involve movement of nutrients and energy through a
30 community as organic matter. The more nutrients available the more energy flows.
31 Hydrological, gaseous and sedimentary cycles are involved. -Water is the medium by which
August 1991
10-3
DRAFT-DO NOT QUOTE OR CITE
-------
1 nutrients make their never-ending odyssey through an ecosystem (Smith, 1980). In gaseous
2 cycles which include carbon, oxygen and nitrogen, the atmosphere is the primary reservoir
3 and in sedimentary cycles, phosphorus, sulfur, calcium, magnesium, and potassium move
4 from the land to the sea and back.
5 Vegetation, through the process of photosynthesis plays a very important role in energy
6 and nutrient transfer. Plants accumulate, use and store carbon, the basic building blocks of
7 large organic molecules, to maintain physiological processes and to form their structure.
8 During photosynthesis, plants utilize energy from sunlight to convert carbon dioxide (CO2)
9 from the atmosphere and water from the soil into carbohydrates. Carbohydrates serve as the
10 raw material for further biochemical synthesis (Waring and Schlesinger, 1985).
11 The energy accumulated and stored by vegetation also is available to other organisms
12 such as herbivores, carnivores and decomposers. Energy and nutrients move from organism
13 to organism in food chains or food webs that become more complex as ecosystem diversity
14 increases (Odum, 1989). Its flow through the biological food chains is unidirectional.
15 Ultimately, it is dissipated into the atmosphere as heat and must be replaced (Barbour et al.,
16 1980; Billings, 1978; Odum, 1989). Nutrients and water can be recycled, fed back into the
17 system, and used over and over again (Barbour et al., 1980; Odum, 1989). The plant
18 processes of photosynthesis, nutrient uptake, respiration, translocation, carbon allocation, and
19 biosynthesis are directly related to the ecosystem functions of energy flow and nutrient
20 cycling. Reduction in diversity and structure in ecosystems shortens the food chains, reduces
21 the total nutrient inventory and returns the ecosystem to a simpler successional stage
22 (Woodwell, 1970).
23
24 10.1.2.3 Ecosystem Response: Impairment of Functions, Changes in Structure
25 Ecosystems respond to stresses through their constituent organisms. In plant
26 communities, individual species differ appreciably in their sensitivity to stresses; the changes
27 that occur within plant communities reflect such differences. The response of plant
28 populations or species to environmental perturbation depends upon their genetic constitution
29 (genotype), life cycles, and the .microhabitats in which the plants are growing. Stresses such
30 as changes in the physical or chemical environment of plant populations apply new selection
31 pressures on individual organisms (Treshow, 1980). A common response in a community
August 1991
10-4
DRAFT-DO NOT QUOTE OR CITE
-------
1 under stress is the elimination of the more sensitive populations and an increase in abundance
2 of species that tolerate or are favored by the stress (Woodwell, 1970; Guderian et al.,. 1985).
3 Factors that influence the rate or amount of energy flow or of nutrient cycling alter the
4 relationships that exist between organisms and their non-living environment. Air pollutants,
5 for example, that limit carbon fixation will shift allocation to new leaves, while factors that
6 limit the availability of nitrogen or water will shift allocation to the roots (Winner and
7 Atkinson, 1986). Such subtle and indirect effects of pollutant exposures, by inhibiting or
8 altering plant physiological processes, decrease the ability of organisms to compete.
9 Increasing pollutant stresses provide selective forces that favor some genotypes, suppress
10 others, and eliminate those species that lack sufficient genetic diversity to survive. Removal
11 of these organisms from an ecosystem can impair ecosystem functions and set the stage for
12 changes in community structure that possibly may have irreversible consequences (Guderian
13 and Kueppers, 1988).
14 Abundant evidence exists to show that plant communities undergo structural changes
15 that reduce biological variation when resistant species become dominant (Miller, 1973, Smith,
16 1980; Treshow, 1980; Woodwell, 1970). In forest communities the selective removal of the
17 larger overstory plants in favor of plants of small stature results in a shift from a complex
18 forest community to the less complex hardy shrub and herb communities (Woodwell, 1970;
19 Miller, 1973). Thus, there is a change in the occurrence, size, and distribution of plants, in
20 species interactions, and in community composition and the processes of energy flow and
21 nutrient cycling are altered. Ultimately, the basic structure of the ecosystem is also changed.
22 Predicting the effects of nitrogen compounds from anthropogenic sources on natural
23 ecosystems involves uncertainties because: (1) it is difficult to determine accurately the
24 atmospheric nitrogen deposition; (2) less is known concerning the response of nonagricultural
25 plant communities to increased supplies of fixed nitrogen than for agricultural crops; and
26 finally, (3) the effects of nitrogen saturation have been studied for only a short time.
27 The next section outlines the nitrogen cycle and mentions changes in the cycle that may
28 result from the increasing additions of nitrogen. The subsequent sections discuss the observed
29 effects of increased nitrogen deposition on terrestrial, wetland and aquatic ecosystems and the
30 possible changes in the nitrogen cycle that may result.
31
August 1991
10-5
DRAFT-DO NOT QUOTE OR CITE
-------
J 10.1.3 The Nitrogen Cycle
2 Nitrogen, one of the main constituents of the protein molecules essential to all life, is
3 recycled within ecosystems. Most organisms cannot use the molecular nitrogen found in the
4 earth's atmosphere. It must be transformed by terrestrial and aquatic microorganisms into a
5 form usable by other organisms. The transformations of nitrogen as it moves through an
6 ecosystem is referred to as the "nitrogen cycle" (National Research Council, 1978). Mature
7 natural ecosystems are essentially self-sufficient and independent of external additions.
8 Modern technology by either adding or removing nitrogen from an ecosystems may be
9 upsetting the relationships that exist among the various components and thus changing its
10 structure and functioning.
11 Nitrogen usually enters plants through the roots by: (1) absorption of ammonia and
12 ammonium, (2) absorption of nitrate (and nitrite), and (3) nitrogen fixation by symbiotic
13 organisms. Therefore, a pollutant that can be converted chemically or biologically into
14 ammonia, nitrate, or nitrite can be used by plants. Nitrogen oxides which fall upon soil have
15 the potential for conversion and adsorption by microbial or chemical action and can enter
16 plants easily through the soil/root interface. Soil-deposited nitrogen, however, can overload
17 the soil/plant system (see below). Gaseous NOX which enters through the leaves can also be
18 converted for plant use since most leaves have enzyme systems which can handle the
19 compounds derived from NOX (see Chapter 9).
20 The term "nitrogen cycle" (Figure 10-1) is used to refer to the transformations of
21 nitrogen as it moves through the environment. In general outline, the nitrogen cycle is
22 identical in terrestrial, fresh water and oceanic habitats; only the microorganisms which
23 mediate the various transformations are different (Alexander, 1977). In terrestrial and aquatic
24 ecosystems, the major nonbiological processes of the nitrogen cycle involve phase
25 transformations rather than chemical reactions. These transformations include
26 (1) volatilization of gaseous nitrogen forms (e.g., ammonia); (2) sedimentation of particulate
27 forms of inorganic nitrogen; and (3) sorption (e.g., of ammonium ions by clays) (National
28 Research Council, 1978). In general, the steps in the nitrogen cycle are as follows:
29 (1) Nitrogen Fixation, (2) Assimilation, (3) Ammonification, (4) Nitrification,
August 1991
10-6
DRAFT-DO NOT QUOTE OR CITE
-------
00
t>
o\
August 1991
10-7
DRAFT-DO NOT QUOTE OR CITE
-------
1 (5) Denitrification. These biological transformations involved in the nitrogen cycle will be
2 discussed below.
3 Under natural conditions, nitrogen is added to ecosystems by fixation of atmospheric
4 nitrogen, deposition in rain, from windblown aerosols containing both organic and inorganic
5 nitrogen, and from the absorption of atmospheric ammonia by plants and soil (Smith, 1980).
6 Nitrogen fixation, the conversion of molecular nitrogen into a biologically available form, is
7 mediated almost entirely by microorganisms in both terrestrial and aquatic habitats
8 (Alexander, 1977).
9 Plants can utilize nitrogen in the form of ammonium or nitrate with equal efficiency and
10 either form can be converted by plants into amino acids, protein and nucleic acids. The
11 organic nitrogen in plants is transferred to herbivores when they eat plants. Herbivores may
12 in turn be eaten and the nitrogen utilized by their predators. The urea and excreta of animals
13 and the organic remains of dead plants and animals are eventually decomposed by
14 microorganisms and transformed into ammonia. Ammonia gas may be (1) volatilized into the
15 atmosphere, (2) converted into nitrates by bacteria, (3) absorbed by plants, or (4) leached into
16 streams, lakes or eventually the ocean where it available for use in aquatic ecosystems.
17 Modern technology is perturbing the cycle by altering the amounts and fluxes of
18 nitrogen in the various portions of the cycle. For example, increased nitrogen oxide
19 emissions from transportation and stationary fossil fuel burning sources over the past 50 years
20 have increased the wet and dry deposition of nitrates and the amount of nitrogen moving
21 through terrestrial and aquatic nitrogen cycles. Crops can utilize only a proportion of the
22 nitrogen fertilizers (containing nitrates, ammonium salts, anhydrous or liquid ammonia, or
23 urea) added to the agricultural soils; leaching and runoff results (Sprent, 1987). Also,
24 ammonia emissions from livestock feedlots have increased the nitrogen moving through the
25 nitrogen cycle. Harvesting of crops, on the other hand, removes nitrogen from
26 agroecosystems and makes them dependent on the addition of inorganic nitrogen fertilizers
27 (Bolin and Arrhenius, 1977). Timber harvesting also removes nitrogen and disrupts the soil-
28 plant-microorganism relationships. Forest clear-cutting increases the loss of nitrates in soil
29 water (Bowden and Bormann, 1986). Burning of the residues left after timber removal may
30 lead to further nitrogen loss (Vitousek, 1981).
31
August 1991
10-8
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
10.1.3.1 Biological Nitrogen Fixation
Nitrogen fixation, the conversion of molecular nitrogen gas (N^ to ammonium
(NH4+), is accomplished by a limited number of free-living and symbiotic (living in the
roots of leguminous plants) bacteria. The ammonia (NH3) formed is available to plants and
other microorganisms. Nitrogen fixation is essential in the maintenance of soil fertility in
terrestrial, aquatic and agricultural ecosystems.
»
10.1.3.2 Assimilation
Plants assimilate inorganic nitrogen from the soil and convert it into organic nitrogen.
All plants, except certain bog and wetland species, are able to assimilate inorganic nitrogen as
either ammonium or nitrate and to convert them into organic molecules such as amino acids,
proteins, and nucleic acids. Bacteria are also important assimilators of inorganic nitrogen in
the soil while algae are the predominant assimilators of inorganic nitrogen in aquatic habitats.
Most plants utilize ammonium more readily than nitrate; however, if no other factors limit
microbial growth, microorganisms will scavenge the available ammonium making it
unavailable. Under these circumstances, nitrate becomes the most important source of
nitrogen for plants (Rosswall, 1981).
10.1.3.3 Ammonification (Mineralization)
Bacteria and fungi form ammonium during the decomposition of dead plants and
animals. Proteins in dead plants and animals, as well as the excretion products of animals,
are decomposed to amino acids. The nitrogen in amino acids in turn are converted into
ammonium. The ammonium may be (1) assimilated by terrestrial or aquatic plants and
microorganism, (2) bound by clay particles in the soil, or (3) converted into nitrates by
microorganisms during nitrification. Ammonification is important in renewing the limited
supply of inorganic nitrogen utilizable by plants.
During ammonification, gaseous ammonia (NH3) may escape into the atmosphere
during the process. Its volatilization is a purely physical process whereby ammonia, in
equilibrium with ammonium (NH4+) in solution, is lost as a gas. Gaseous losses are
significant if pH is below 7.5 (Reddy and Patrick, 1984). Ammonia volatilization can be
mediated by biological activity to the extent that organisms can alter the pH of their
August 1991
10-9
DRAFT-DO NOT QUOTE OR CITE
-------
1 environment. Ammonia losses from wetlands are normally significant because submerged
2 and wetland soils generally have pH values between 5.0 and 7.2 (Ponnanperuma, 1972).
3
4 10.1.3.4 Nitrification
5 Nitrification is the two-step process during which microorganisms first convert
6 ammonium (NH4+) to nitrite (NO2") and then to nitrate (NO3~). In the first step, several
7 genera of bacteria including the genus, Nitrosomonas) reduce ammonium to nitrite. The
8 second step is accomplished by several genera of bacteria including Nitrobacter that reduced
9 nitrite to nitrate (Reddy and Patrick, 1984; Atlas and Bartha, 1981). Nitrification is strictly
10 an aerobic process and only oxygen can serve as the electron acceptor. Nitrification can
11 occur in manure piles, during sewage processing, in soil and in marine environments in the
12 oxygenated water column above the anaerobic sediments or within the surface of oxidized
13 layers of sediments. Recent studies suggest that nitrous oxide (N2O) is produced during
14 nitrification. Bowden (1986) points out, however, that in the field nitrous oxide production
15 via nitrification is controlled by the oxidation status of the soil.
16 Other than atmospheric transformations of NOX to nitrates, nitrification is the sole
17 natural source of nitrate in the biosphere (National Research Council, 1978). Nitrate is the
18 predominant nitrogenous ion in precipitation (U.S. Environmental Protection Agency, 1982).
19 It is at this stage that the nitrogen cycle has been most influenced through agricultural
20 practices (Delwiche, 1977; Bolin and Arrhenius, 1977). Natural processes are unable to
21 produce sufficient nitrogen to grow the crops needed to feed humanity. This has led to the
22 development and increasing use of industrially made fertilizers. • In 1970, Delwiche (1970)
23 estimated that the amount of nitrogen fixed annually since 1950 for the production of
24 fertilizer equaled the amount that was fixed by all terrestrial ecosystems before the advent of
25 modern agriculture.
26 Nitrates, whether added to the soil (1) as fertilizers, (2) by nitrification, or (3) from
27 atmospheric deposition, may:
28 • be utilized by microorganisms,
29 • be taken up by plants,
30 • be lost through surface runoff into streams, rivers, lakes, wetlands or
31 oceans,
32 • percolate into the ground water, or
33 • escape as gas to the atmosphere (Buckman and Brady, 1969).
August 1991
10-10
DRAFT-DO NOT QUOTE OR CITE
-------
1 10.1.3.5 Denitrification
2 Denitrification is an anaerobic bacterial process during which nitrates are converted into
3 atmospheric nitrogen gas. Nitrates (NO3~), are converted into nitrites (NO2~)> then gaseous
4 nitrous oxide (N2O) and finally into nitrogen gas (N^, which escapes into the atmosphere.
5 Under acidic conditions in the soil, nitrites rarely accumulate but are spontaneously
6 decomposed into nitric oxide (NO). Under alkaline conditions, they are biologically
7 converted into nitrous oxide and molecular nitrogen (Alexander, 1977).
8 Through denitrification, nitrogen becomes unavailable to most plants and
9 microorganisms because it enters the large atmospheric reservoir where its residence time
10 may be as long as 107 years (Delwiche, 1977). Nitrous oxide has a much shorter residence
11 time. The photochemical decomposition of nitrous oxides is the main stratospheric source of
12 nitrogen oxides (Delwiche, 1977). Nitrous oxide has been implicated in global warming
13 (BoUe.-et.al., 1986). ;.
14 Nitrogen resides in five major reservoirs: (1) primary rocks, (2) sedimentary rocks,
15 (3) deep-sea sediment, (4) the atmosphere, and (5) the soil-water pool. The web of pathways
16 and fluxes by which oxides of nitrogen are produced, transformed, transported and stored in
17 the principal nitrogen reservoirs are commonly referred to as the nitrogen cycle are outlined
18 above. An understanding of the nitrogen cycle is important in placing in perspective human
19 intervention as discussed in other sections of this chapter.
20
21
22 10.2 DRY DEPOSITION RATES OF REACTIVE "N" FORMS
23 Deposition processes result in the removal of reactive nitrogen compounds from the
24 atmosphere, and their subsequent deposition onto landscape surfaces (e.g., foliage, bark,
25 soil). The fate of dry deposited compounds can be either adsorption to surfaces or absorption
26 (i.e., uptake or incorporation) by surfaces. By quantifying the link between atmospheric
27 processes and deposition of pollutants to plants, deposition measurements provide valuable
28 input data for models of atmospheric chemistry, biogeochemical cycling, and may help
29 explain how pollutants affect plants (Baldocchi et al., 1987, 1988; Hosker and Lindberg,
30 1982; Taylor etal., 1988).
August 1991
10-11
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
Dry deposition characteristics of nitrogen dioxide (NO^, nitric oxide (NO), nitric acid
vapor (HNO3), ammonia (NH3), and particle forms (NO3~ and NH4+) have been reported in
the literature and are discussed in the following sections. Ammonia is not an oxide of
nitrogen, but when present at high concentrations in the atmosphere, it contributes to the total
amount of N deposited on landscape surfaces, and by dissolving in aerosols NH3 may
enhance HNO3 removal in wet precipitation (Erisman et al., 1988). Therefore, ammonia
deposition data is included here. Deposition data is unavailable for other potentially
important reactive forms of N: nitrous acid (HNO2), dinitrogen pentoxide (N2O5), and the
gaseous nitrate radical (NO3). Pernitrate species, such as peroxyacetyl nitrate (PAN), will
not be discussed because they are described in another Air Quality Criteria Document (U.S.
Environmental Protection Agency, 1986). Nitrous oxide (N2O), the most abundant nitrogen
oxide, will not be discussed because it is virtually inert in the troposphere and shows no
tendency for deposition (Singh, 1987).
Garner et al. (1989) summarized available information on ambient air concentrations for
nitrogen oxides and made the following conclusions:
1. Nitrogen oxides are rarely if ever found in concentrations sufficient to
cause visible injury to vegetation.
2. In high elevation forests typically away from urban sources of pollution,
concentrations of nitrogen oxides are usually below or at the detection
limits of available monitoring equipment (concentrations range from
<0.003 to occasional peaks of 0.05 ppm).
3. In near-urban or rural forests, concentrations seldom exceed 0.010 ppm
(overall range from <0.005 to 0.3 ppm).
4. In urban areas of eastern North America annual average nitrogen oxide
concentrations are around 0.07 ppm with values ranging from < 0.005
to 0.4 ppm.
A number of recent studies in remote areas have shown that air concentrations of NO, NO2,
and HNO3 are commonly less than 0.005 ppm with HNO3 concentrations typically being
lower (Cadle et al., 1982; Fahey et al., 1986; Kelly et al., 1984; Lefohn and Tingey, 1984).
In rural areas closer to sources of urban pollution, NO2 and HNO3 concentrations have been
measured in the 0.010-0.030 ppm and 0.001-0.003 ppm ranges, respectively (Bytnerowicz
August 1991
10-12
DRAFT-DO NOT QUOTE OR CITE
-------
1 et al., 1987; Kelly et'al., 1989; Lefohn and Tingey, 1984). A detailed summary of current
2 information on the air chemistry and concentrations of reactive nitrogen compounds can be
3 found in Chapters 5 and 7 of this document.
4 There are several general review articles for additional information ori the deposition of
5 N forms to vegetation and other landscape surfaces. Hosker and Lindberg (1982) discuss
6 factors controlling pollutant deposition and capabilities for predicting interactions between
7 atmospheric substances and vegetation. McMahon and Denison (1979) provide a more
8 extensive summary of particle deposition. Sehmel (1980) summarizes particle and gas dry
9 deposition for a wide range of depositing materials. Taylor-et'al. (1988) review pollutant
10 deposition to individual leaves and plant canopies with particular emphasis on physiological
11 sites of regulation. World Health Organization (1987) also provides an extended discussion
12 of deposition of N forms important to the establishment of air quality guidelines.
13
14 10.2.1 Types of Measurements
15 Dry deposition measurements have been conducted in the field at the forest canopy level
16 or in chambers using individual plant leaves (van Aalst and Diederen, 1985). Canopy level
17 measurements are based on the assumption that deposition is a vertical flux from the
18 atmosphere to a defined landscape area restricted by a series of pathway resistances. Leaf-
19 level measurements in chambers, which ignore the atmospheric transport process by inducing
20 turbulent mixing above the surface of leaves, also assume a series of resistances to pollutant
21 gas deposition. Leaf-level and canopy measurements are normalized to leaf and ground
22 areas, respectively. Although not yet applied to pollutant gas deposition, Matson and Harriss
23 (1988) have suggested the use of aircraft-based measurements to study gas exchange over
24 wide landscape areas.
25 Canopy measurements typically employ either the eddy correlation or the flux gradient
26 micrometeorological techniques. Both techniques require that measurements be conducted
27 under ideal conditions (e.g., flat, homogeneous, and extensive landscape area), but some
28 progress in applying these techniques to more complex terrain has been made (McMillen,
29 1988; Hicks et al., 1984). The eddy correlation technique measures vertical, turbulent flux
30 directly from calculations of the mean covariance between wind velocity and pollutant
31 concentration (Wesely et al., 1982). The flux gradient or "profile" technique estimates
August 1991
10-13
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
S
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
vertical flux from a concentration profile and eddy exchange coefficients (Erisman et al.,
1988; Huebert et al., 1988). One of the most difficult problems with dry deposition
estimates of N species, based on micrometeorological methods, stems from the inability to
measure the appropriate atmospheric concentrations. Homogeneous gas phase reactions and
gas/particle interactions of HNO3 and NH3 (Appel and ToMwa, 1981), and interferences of
HNO3 with some NOX sensors (van Aalst and Diederen, 1985) are two examples of the ;
problems often encountered. Many N species are so reactive in the canopy air space that
their concentrations change significantly during the course of micrometeorological
measurements, resulting in misleading flux data (Hicks et al., 1989). Businger (1986) and
Baldocchi (1988) provide more extensive discussion of the benefits and/or pitfalls of the
canopy measurement techniques.
Comparisons between throughfall or precipitation nitrate and ammonium ion
concentrations have also been used to calculate particulate N deposition to forest canopies
(Gravenhorst et al., 1983; Lovett and Lindberg, 1984). However, the reactivity of trace N
gases, their absorption by foliar surfaces (Norby et al., 1989; Garten and Hanson, 1990), and
the technique's inability to distinguish gaseous from particle forms (e.g., NO3" vs. HNO3)
may lead to large errors.
Three techniques have been used for leaf-level measurements. The most common
approach is based on mass-balance principles in which the leaf surface is enclosed in an
environmentally controlled chamber and pollutant concentrations are compared at the inlet
and outlet (Jarvis et al., 1971). The mass-balance technique can be applied to individual
leaves and branches (Rogers et al., 1977; Rowland-Bamford and Drew, 1988) or to enclosed
crop canopies (Bennett and Hill, 1973; Hill, 1971). Less commonly, isotopic labeling of the
exposure gas with 15N has been used to evaluate rates of deposition (Okano et al., 1988;
Vose and Swank, 1990). Leaf washing techniques compare extracts from leaves exposed to
pollutants and appropriate controls. The difference in ion concentrations between treated and
control wash solutions is used to calculate rates of deposition (John et al., 1985; Dasch,
1987). Leaf-wash techniques may underestimate deposition because absorption or
translocation processes remove pollutants from the leaf surface (Taylor et al., 1988; Garten
and Hanson, 1990). Further, the leaf-wash method can not distinguish various sources of
nitrate deposited as HNO3, NO3, or particulate NO3" (Dasch, 1987).
August 1991
10-14 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
10.2.2 Expressions of Deposition
Rates of pollutant deposition determined from canopy or leaf level measurements can be
expressed with similar equations. The rate of deposition of pollutant gases to a canopy
surface has been defined as:
* (Gz - C0>
(1)
where Fc is flux to the canopy in nmol nf2 s"1, Cz is the concentration at the height of the
o
measurement (nmol m"3), C0 is the concentration at receptor sites in the canopy (nmol m" ),
and Vd is the overall deposition velocity in units of m s-1. The Vd is the reciprocal of the
total canopy resistance to flux. An analogous equation can be derived for leaf-level, chamber
measurements: . ' - -' •>
(Ca --
(2)
where Fl is flux to leaves, Ca is the concentration of pollutant in the air around the leaf, C{
is the concentration of pollutant in or on the leaf (often equal to 0), and Kt is the
conductance of the leaf to pollutant gas transfer.
Both Vd and Kx represent concentration corrected deposition rates, and they are the :
standard variables used to compare deposition, characteristics of pollutant gases and receptor
surfaces. Although Vd and Kt have the same units, they are based on different receptor areas
and characterize processes at different scales of resolution. Therefore, the following
conversion has been suggested as a first approximation for scaling between canopy and leaf
measurements of pollutant deposition so that data obtained with either technique can be
compared: .
...,yd =,K! *LAI .-. .• (3)
where LAI is the leaf area index of the canopy appropriate to the Vd variable (Dasch, 1987;
Dolske, 1988; Hanson etal., 1989; Hicks etal., 1987; Jarvis, 1971; O'Dell et al., 1977).
For a given plant material and defined exposure, Vd should always be larger than Kx when
canopy leaf area index is greater than one. This first-order conversion is admittedly crude
August 1991
10-15
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
but useful. Complex models are required to rigorously scale measured Kj data to application
at the canopy level of resolution (Baldocchi, 1988; Baldocchi et al., 1987; Hicks et al., 1987;
Kramm, 1989) because nonlinear processes are involved and driving variables change with
depth in the canopy.
10.2.3 Processes Governing Deposition of Gases and Particles
Dry deposition of gases and particles to foliar and non-foliar surfaces refers to the
transfer of N species between the free atmosphere and landscape surfaces. Dry deposition
i
processes need to be understood because they represent the first step in the transfer of
pollutants to physiological sites of action in the leaf interior (Taylor et al., 1988) that are
responsible for most deleterious effects on plants. Detailed discussions of the factors
influencing dry deposition of gases and particles have been published (Hosker and Lindberg,
1982; Sehmel, 1980; Taylor et al.> 1988). The reader is also directed to Section 10.2.4 for
additional discussion of reactive nitrogen gas deposition to leaves and leaf interior spaces.
Pollutant gas deposition to plant surfaces is controlled by atmospheric turbulence,
physical and/or chemical properties of gases, the presence of a chemical potential gradient
between the atmosphere and receptor sites, and the nature and activity of plant surfaces
(Table 10-1). Hosker and Lindberg (1982) divided gaseous pollutant compounds into three
groups based on the processes governing their deposition and assigned reactive nitrogen
compounds to each group as shown below.
(1) compounds able to adsorb readily to all surfaces [HNO3, NH3].
(2) compounds that interact with leaves primarily after diffusion through
stomata into interior leaf air spaces [NO2 and to some extent NH3].
(3) compounds that exchange slowly with plants independent of the pathway
for deposition [NO, N2O].
Recent data (Kisser-Priesack et al., 1987) suggest that NO2 and NO are also deposited onto
and through the cuticle; a feature appropriate to Hosker and Lindberg's Category #1
i
compounds.
August 1991
10-16 DRAPT-DO NOT QUOTE OR CITE
-------
TABLE 10-1. FACTORS INFLUENCING DRY DEPOSITION
OF REACTWE NITROGEN COMPOUNDS
(Modified after Sehmel, 1980)
Chemical Properties of Depositing Material
Micrometeorological Variables
Aerodynamic resistance:
-mass transfer
-heat
-momentum
-I/deposition velocity
I
Diffusion effect of:
-canopy structure
-extent of fetch
Particles
Particle size:
-diameter
^density
-agglomeration
Diffusion:
-Brownian
-eddy
-eddy
Impaction
•• Qases
Partial Pressure
-solubility •
-concentration
Chemical activity/
reactions
, ' ,
Diffusion:
,•
-molecular
Receptor
Surface Variables
Abiotic features
Accommodation:
-dew
-exudates
-wax
-pubescence
Reactive sites:
-area
-prior loading
-adsorption
-absorption
Friction velocity
Surface roughness length
Zero plane displacement
Wind velocity
Turbulence
Temperature
Relative humidity
Precipitation
Solar radiation
Gravitational settling
Electrostatic effects
Biotic features
Stomatal
-conductance
-diurnal pattern
Plant metabolic rate
-assimilation
-cell pH
1
2
3
4
5
The theory of particle deposition has been described and discussed in depth in several
recent papers (Davidson and Wu, 1990; McMahon and Denison, 1979; -Nicholson, 1988;
Sehmel, 1980). These authors propose three characteristic features of dry particle deposition:
August 1991
10-17 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
(1) Particles greater than 10 jttm exhibit a variable deposition velocity (Vd)
between 5 and 110 mm s"1 dependent on frictional velocities, while a
minimum particle deposition velocity has been shown for particles in the
size range 0.1-1 #m (Figure 10-2).
(2) Deposition velocity of particles (Vd) is approximately a linear function of
friction velocity.
(3) Deposition of particles between the atmosphere and a forest canopy is
from 2 to 16 times greater than deposition in adjacent open terrain (i.e,
grasslands or other vegetation of low stature).
Theoretically based models for predicting particle deposition velocities have recently
been published by Bache (1979a, 1979b), Davidson and Wu (1990), and Noll and Fang
(1989). Dolske (1988) claims that dry deposition, whether in the form of gases or particles,
has from 3 to 20 times the potential of wet deposition to modify the chemical
microenvironment of foliar surfaces. This claim was made based on the "cyclic reactivation"
of dry deposition by dew and rain which appears to, dissolve and mobilize, but not necessarily
remove the pollutants from the foliar surface.
Independent of the site of deposition of gases or particles (internal vs. cuticular) the
concentration of the pollutant in ambient air is representative of the driving force responsible
for direct and indirect effects on plant physiological processes. However, because the
chemical nature of all pollutants are not the same a single "time-averaged" concentration
(e.g., 24 h vs. daylight means) might not be appropriate in all cases. For example, a 24 h
mean concentration is appropriate for the largely cuticular deposition observed for aerosol
particles and HNO3, but a daylight mean would be better for those pollutant gases whose
deposition is tightly controlled by stomatal aperture limitations to diffusion (e.g., NO, NO2).
10.2.4 Deposition of "N" Forms to Foliar Surfaces
Reported deposition velocities or conductances for NO2, NO, HNO3, NH3, and
particulate nitrogen forms are presented in Tables 10-2 through 10-10. Each table is
organized by plant species or deposition surface and, unless noted otherwise, the listed
deposition velocities correspond to daytime conditions. Actual Vd values are highly variable
reaching maximum and minimum values during midday and night periods, respectively. Two
types of tables are used to present the data for each of the four gases: tables covering
August 1991
10-18
DRAFT-DO NOT QUOTE OR CITE
-------
No resistance below and
atmospheric diffusion from
1 cm to 1 m
brownlan below and
pherio diffusion above
Indicated height
10'1 1
Particle Diameter, urn
Figure 10-2. Predicted deposition velocities (Vd) at 1 meter for a friction velocity of
30 cm s"1 and particle densities of 1,4, and 11.5 g cm"3 (Sehmel, 1980).
August 1991
10-19
DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-2. CONDUCTANCE (K^ OF NO2 TO LEAF SURFACES
Species
Austrian pine (Pinus nigra)
Barley (Hordeum vitlgare)
Bean (Phaseolus vulgaris)
g, = 0.26
g. = 0.05
Chinese hibiscus
(Hibiscus rosa-sinensis)
Cucumber (Cucmnis salivas')
Diffenbachia maculata
Douglas fir (Pseudotsuga
mensiesif)
[Mirb.] Franco
English ivy (Hedera helix)
European White Birch
(Belula pendula)
Ficus benjamina
Hedera canariensis
Honey locust
(Gleditsia triancanthos')
Indian rubber
(Ficus elastica)
Loblolly pine (Pinus taeda)
Concentration
ppmv (/ig/m"3)a
0.400
0.3
0.3
0.04
0.16
0.5
1
3
7
1.0
1.0
4.0
4.0
0.500
1.0
4.0
0.400
1.0
4.0
<0.06
0.400
0.400
1.0
4.0
1.0
4.0
0.400
1.0
4.0
0.020
rv *ibc
L-^-i J
rirm s"1
0.3d
0.5
0.5
0.7
0.1
1.0
0.8
0.85
0.63
0.69
0.79
0.54
0.65
1.1
0.49
0.31
0.2d
0.56
0.29
3.2
0.2
0.1
0.47
0.19
0.62
0.35
0.2
0.86
0.69
0.6
Method Reference
Chamber Elkiey et al. (1982)
Chamber Rowland-Bamford and Drew (1988)
15N Rowland-Bamford and Drew (1988)
Chamber Fuhrer and Erismann (1980)
Chamber Fuhrer and Erismann (1980)
15N Okano et al. (1988)
Chamber Srivastava et al. (1975)
Chamber Srivastava et al. (1975)
Chamber Srivastava et al. (1975)
Chamber Saxe (1986)
Chamber Saxe (1986)
Chamber Saxe (1986)
Chamber Saxe (1986)
15N Okano et al. (1988)
Chamber Saxe (1986)
Chamber Saxe (1986)
Chamber Elkiey et al. (1982)
Chamber Saxe (1986)
Chamber Saxe (1986)
NAe Freer-Smith (1983)
Chamber Elkiey et al. (1982)
Chamber Elkiey et al. (1982)
Chamber Saxe (1986)
Chamber Saxe (1986)
Chamber Saxe (1986)
Chamber Saxe (1986)
Chamber Elkiey et al. (1982)
Chamber Saxe (1986)
Chamber Saxe (1986)
Chamber Hanson et al. (1989)
August 1991
10-20
DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-2 (cont'd). CONDUCTANCE (K^ OF NO2 TO LEAF SURFACES
Species
Lombardy poplar
(Populus nigra)
Maize (Zea mays)
Mountain ash (Sorbus aria)
Nephrolepsis exaltala
Norway spruce (Picea abies)
Petunia (Petunia hybrida)
Prunus sargentii
Radish (Raphanus sativus)
Red maple (Acer rubrum)
Red spruce (Picea rubens)
Spruce (Picea sp.)
dormant
Scots pine (Pinus sylvestris)
current shoot
day
night
1-year shoot
day
night
2-year shoot
day
night
branches
branches
branches
dormant
dormant (field)
dormant (field)
dormant (field)
dormant (lab)
dormant (lab)
dormant (lab)
Concentration
ppmv Oig/m~3)a
<0.06
0.2
0.5
0.5
1.0
0.400
1.0
4.0
0.400
0.400
0.400
0.500
0.020
0.020
0.006-0.03
NA
NA
NA
NA
0.093 (175)
0.093 (175)
0.001
0.005-0.01
0.02-0.03
0.106 (200)
0.026 (50)
0.066 (125)
0.119 (225)
0.053 (100)
0.159 (300)
0.265 (500)
mm s"1
2.9
0.6
0.8
0.9
0.7
0.2
0.48
0.22
0.2d
0.6
0.1
1.9
1.8
0.4
«0.3
2.2-7.9
0.6-6.0
10
3.8
10.6
5.8
-l.l-2.1f
0.9-1.7
1.2-3.5
<1.0
0.8
0.6
0.6
0.2
0.2
0.2
Method
NA
15N
15N
15N
15N
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
15N
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Reference
Freer-Smith (1983)
Okano et al. (1986)
Okano et al. (1986)
Okano et al. (1986)
Okano et al. (1986)
Elkiey et al. (1982)
Saxe (1986)
Saxe (1986)
Elkiey et al. (1982)
Elkiey and Ormrod (1981)
Elkiey et al. (1982)
Okano et al. (1988)
Hanson et al. (1989)
Hanson et al. (1989)
Granat and Johansson (1983)
Grennfelt et al. (1983)
Grennfelt et al. (1983)
Grennfelt et al. (1983)
Grennfelt et al. (1983)
Grennfelt et al. (1983)
Grennfelt et al. (1983)
Johansson (1987)
Johansson (1987)
Johansson (1987)
Grennfelt et al. (1983)
Skarby et al. (1981)
Skarby et al. (1981)
Skarby et al. (1981)
Skarby et al. (1981)
Skarby et al. (1981)
Skarby et al. (1981)
August 1991
10-21 DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-2 (cont'd). CONDUCTANCE (Kj)' OF NO2 TO LEAF SURFACES
Concentration [Kj]130
Species ppmv (/ig/m"3)a mm s"1
Sunflower
(Helianthus annus)
Sweet pepper
(Capsicum annum)
Sycamore maple
(A. platanoides)
Sycamore
(Platantis occidentalis)
Sorghum (Sorghum vulgare)
Tobacco (Nicotiana
tabacum)
Tomato
(Lycopersicon esculentum)
light
dark
White ash
(Fraxinus americand)
White oak (Quercus alba)
White fir (Abies concolor)
White pine (Pinus strobus)
Yellow-Poplar
(Liriodendron tulipiferd)
0.2
0.3
0.5
0.5
1.0
2,0
1.5
NA
0.400
0.020
0.500
0.500
0.500
1.5
1.5
0.020
0.020
0,400
0.020
0.020
1.1
3.0
2.3
2.2
2.1
3.4
0.02-1.6
1.3
0.1
4.1
0.6
1.3
1.5
2.0-2.8
1.1-1.6
0.7
1.3
0.3d
0.4
1.5
Method
1 I5N
15N
15N
I5N
15N
15N
Chamber
NA
Chamber
Chamber
15N
15N
15N
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Reference
Okano et al. (1986)
Okano and Totsuka (1986)
Okano et al. (1986)
Okano et al. (1986)
Okano et al. (1986)
Okano and Totsuka (1986)
Rowland et al. (1985)
Law and Mansfield (1982)
Elkiey et al. (1982)
Hanson et al. (1989)
Okano et al. (1988)
Okano et al. (1988)
Okano et al. (1988)
Murray (1984)
Murray (1984)
Hanson et al. (1989)
Hanson et al. (1989)
Elkiey et al. (1982)
Hanson et al. (1989)
Hanson et al. (1989)
Tor NO2 at 25 °C [1 /tg/m3 = 0.000531 ppmv].
"Data are presented as a range or the mean of reported values.
°Data for broadleaved plants and conifers are presented on a one-sided and total leaf area basis, respectively.
dBased on a one-sided leaf area.
*NA = not available.
'Negative values represent evolution of NO2 from leaves.
1 leaf-level or canopy-level measurements. If a cited paper lumped data for NO and NO2
2 together as NOX, that data is presented in Table 10-2 along with the information on NO2, but
3 it is indicated as being for NOX. If the original authors did not calculate Kt or Vd,
August 1991
10-22
DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-3. DEPOSITION VELOCITY (Vd) OF NO2
TO PLANT CANOPY SURFACES
Concentration •
Species ppmv (/ig/ni3)a
Alfalfa (Medicago sativa)
day
night
Grass
lawn(NOx)
pasture (NOJ
Oats (Avena sativa)
day
night
Soybean (Gfycine max [L.J Merr.)
day
night
Spruce (Picea sp.) (NOJ
0.05
0.1
0.24
0.16
0.017 (32.4)
NAd
0.08
0.08
0.08
0.008-0.12
0.008-0.12
0.018
0.029
mm s"1
19
20
10.4
4.1
1.0-3.0
-26-15
12.5
12.5
4.2
3.6
0.7
28
20
Method
Chamber
Chamber
Chamber
Chamber
Flux grad.
Flux grad.
Chamber
Chamber
Chamber
Eddy Corr.
Eddy Corr.
Gradient.
Gradient
Reference
Hill (1971)
Bennett and Hill (1973)
Tingey (1968)
Tingey (1968)
Delany and Davies (1983)
Duyzer et al. (1983)
Hill (1971)
Tingey (1968)
Tingey (1968)
Wesely et al. (1982)
Wesely et al. (1982)
Enders and Teichmann
Enders and Teichmann
(1986)
(1986)
Tor NO2 at 25 °C [1 /*g/m3 = 0.000531 ppmv].
bData are presented as a range or the mean of reported values.
cData are based on ground area under the canopy.
dNA = not available.
1 concentration and flux data from the original papers were used in Equations 1 or 2 to
2 generate the values reported in the following tables.
3
4 10.2.4.1 Nitrogen Dioxide
5 Direct measurements of NO2 deposition to crop species are widely reported (e.g.,
6 Bennett and Hill, 1973; Okano and Totsuka, 1986; Rogers et al., 1979b; Sinn et al., 1984;
7 Wesely et al., 1982), but fewer observations are available for woody plant species (Elkiey
8 et al., 1982; Grennfelt et al., 1983; Rogers et al., 1979) and fewer still for woody plants
9 using near-ambient concentrations of NO2 (Hanson et al., 1989; Johansson, 1987; Skarby
August 1991
10-23 DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-4. CONDUCTANCE (K^ OF NO TO LEAF SURFACES
Species
Chinese hibiscus
(Hibiscus rosa-sinensis)
Diffenbachia maculata
English ivy (Hedera helix)
Ficus benjamina
Hedera canariensis
Indian rubber
(Ficus elasticd)
Nephrolepsis exaltala
Pine/spruce dormant
Scots pine (Pinus sylvestris)
dormant (field)
dormant (lab)
Concentration
ppmv (/*g/m~3)a
4.0
4.0
4.0
1 4.0
4.0
4.0
4.0
0.0005-0.002
variable
0.122 (150)
0.244 (300)
0.407 (500)
[KJta
tmn s"1
0.22
0.34
0.10
0.10
0.13
0.34
0.22
«0.3
«0.1
0.04
0.04
0.05
Method
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber1
Chamber
Chamber
Chamber
Reference
Saxe (1986)
Saxe (1986)
Saxe (1986)
Saxe (1986)
Saxe (1986)
Saxe (1986)
Saxe (1986)
Granat and Johansson (1983)
Johansson (1987)
Skarby et al. (1981)
Skarby et al. (1981)
Skarby et al. (1981)
'For NO at 25 °C 1 /tg/rnT3 = 0.000814 ppmv.
'Data are presented as a mean or range of reported values.
"Data for broadleaved plants and conifers are presented on a one-sided and total leaf area basis, respectively.
TABLE 10-5. DEPOSITION VELOCITY (Vd) OF NO
TO PLANT CANOPY SURFACES
Species
Alfalfa
(Medicago sativa)
Concentration
ppmv (Atg/m"3)a
0.100
0.050
[Vd]bc
mm s"1
1.7
1.0
Method
Chamber
Chamber
Reference
Bennett and Hill (1973)
Hill (1971)
•For NO at 25 °C 1 /ig/m'3 = 0.000814 ppmv.
bData are the mean or a range of reported values.
TData are based on ground area under the canopy.
August 1991
10-24
DRAFT-DO NOT QUOTE OR CITE
-------
TO LEAF SURFACES
Reference
Species
American elm (Ulmus americand)
Austrian pine (Pinus nigra)
Pin oak (Quercus palustris)
Red maple (Acer rubrum)
Red spruce (Picea rubens)
Sycamore (Platanus occidentalis)
White oak (Quercus alba)
White pine (Pinus strobus)
Concentration
ppmv (jug/m~3)a
1.2 - 0.012
0.012-1.2
0.012-1.2
0.02-0.03
0.058-0.067
0.02-0.07
0.04-0.07 ,
37-500 (95-1,288)
[Kj] :n
mm s"1
12
2.0
4.4
3.3°
2.6
1.1
2.2
0.4-0.8
Method
Chamber Das
Chamber Das
Chamber Das
Chamber Hai
Chamber Hai
Chamber Haj
Chamber Ha
Chamber M£
("1C
Hanson et al. (1991)
Hanson et al. (1991)
Hanson et al. (1991)
Hanson et al. (1991)
Marshall and Cadle
(1989)
Tor HN03 at 25 °C 1 /*g/rn3 = 0.000388 ppmv.
"Data for broadleaved plants and conifers are presented on a one-sided and total leaf area basis, respectively.
The data from Hanson et al. (1991) represent cuticular deposition only.
1 et al., 1981). Tables 10-2 and 10-3 provide a comprehensive listing, by plant species, of
2 current data on the deposition of NO2 to leaf and canopy surfaces, respectively. Data are
3 also available for potato plants (Sinn et al., 1984), but conversion of that data to standard
4 units was not possible from the information supplied.
5 Nitrogen dioxide is deposited to plants over a range of concentrations from as little as
6 0.005 ppmv (Johansson, 1987) to those as great as 4 to 7 ppmv (Saxe, 1986; Srivastava
7 et al., 1975). The rate of deposition increases in proportion to rising ambient NO2
8 concentrations (Sinn et al., 1984; Srivastava et al., 1975; Skarby et al., 1981). At low
9 concentrations of NO2 (0.0013 ppmv [2.4 gg nf3]), Johansson (1987) observed no deposition
10 in Scots pine. Johansson suggested that his data indicated a "compensation point" at which
11 rates of NO2 deposition and evolution balance out. The compensation point was reported in
12 the 0.001 to 0.003 ppmv range. If this compensation point is a general phenomenon it would
13 indicate little potential for NO2 deposition at concentrations common across many non-urban
14 , areas of the United States (i.e., areas of NO2 concentration <0.005 ppmv). However, more
15 recent observations have shown that sunflower (Helianthus annuus) does not exhibit an NO2
16 compensation point (Foerstel et al., 1989). Additional discussion of the deposition of NO2
17 into leaves can be found in Section 9.4.1.
August 1991
10-25 DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-7. DEPOSITION VELOCITY (Vd) OF HNO3 TO CANOPY SURFACES
Species
Barley (Hordeuni)
Beets (Beta)
Crop canopies
wind = 1ms'1
wind = 4 m s"1
Forest
wind = 1ms"1
wind = 4 m s"1
Grass (pasture)
wind = 1ms"1
wind = 4 m s"1
Pine (Pinus)
Potato (Solatium)
Spruce (Picea)
Wheat (Triticum)
lj=- »— — — iuijii ijij.il. I-
Concentration
ppmv Og/m"3)"
NA
NAd
0.0001-OJ0005
NA'
NA
0.001-0.002
NA
NA,
NA:
NA
< 0.001 1 (2)
< 0.002 (2.6-4.3)
< 0.002 (3.2)
< 0.003
NA
NA:
NA
NA
0.001
NA
NA ,
NA
======
CKJ"
mm s"1
77
14
5-20
14
50
22-50
20-50
20-60
40
100
40
17-49
25
6
3-18
7-37
5
23
20-70
4
60-120
50-260
=====
Method
Flux grad.
Flux grad.
Model
NA
NA
Flux grad.
Model
Model
NA
NA
Flux grad.
Flux grad.
Flux grad.
Eddy flux.
Flux grad.
Flux grad.
NA
NA
Leaf Wash
Flux grad.
Model
Flux grad.
-
Reference
Harrison et al. (1989)
Harrison et al. (1989)
Meyers and Hicks (1988)
Fowler et al. (1989)
Fowler et al. (1989)
Meyers et al. (1989)
Hicks et al. (1985)
Hicks and Meyers (1988)
Fowler et al. (1989)
Fowler et al. (1989)
Erisman et al. (1988)
Huebert (1983)
Huebert and Robert (1985)
Huebert et al. (1988)
van Aalst & Diederen
(1985)
Harrison et al. (1989)
Fowler et al. (1989)
Fowler et al. (1989)
Dasch (1987)
Harrison et al. (1989)
Hicks et al. (1985)
Dollard et al. (1987)
•For HN03 at 25 °C 1 ^g/m"3 = 0.000388 ppmv.
bData are means or a range of the reported values.
"Data are based on ground area under the canopy.
dNA = not available.
1
2
3
4
5
6
7
Numerous studies have confirmed the control of stomatal aperture on NO2 deposition
using a variety of techniques (Hanson et al., 1989; Rogers et al., 1977; Rogers et al.,
1979a,b; Saxe, 1986; Wesely et. ial., 1982; see also Section 9.4.1). In addition, Murray
(1984), using a tomato mutant whose stomata did not close in the dark, claimed to have
found a direct relationship between light and NO2 deposition.
Until recently it was assumed that cuticular deposition of NO2 was negligible. Recent
studies by Lendzian and Kerstiens (1988) and Kisser-Priesack et al. (1987) clearly
August 1991
10-26 DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-8. CONDUCTANCE (KT) OF NH3 TO LEAF SURFACES
.
Species
Bean (Phaseolus vulgaris)
26.6 °C
33.4 °C
Cotton
(Gossypium hlrsutum)
Fescue
Heather/purple moor grass
(Calluna/Molina)
Italian rygrass
(Lolium multiflorum)
Maize (Zea mays)
Oats (Avena )
Orchard grass
Populus euramericana
Soybean (Glycine max)
Sunflower
(Helianthus annuus)
Tomato
(Lycopersicon esculentum)
Tobacco
(Nicotiana tabacum)
Wheat (Triticum)
Concentration
ppmv (jug/m"3)"
0.002
0.0035
0.005
0.005
0.008
0.14
0.071 (50)
0.144(100)
0.288 (200)
0.502 (350)
0.063 (44)
0.331
0.341
NAd
. 22.6 (16)
735 (520)
0.034 (24)
0.320
0.200
0.283
0.072
0.143
0.037 (26)
0.170
0.045 (31)
0.148
0.173
0.277
mm s"1
0
2
3-11
0
6-32
13
2-5
2-6
2.5-6
2-6
2
7
15
4
3
28
6.5
4
13
10
0.5-5
0.5-9
4
11
4
10
6
15
•— = —
Method
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Estimated
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
.
Reference
Farquhar et al. (1980)
Farquhar et al. (1980)
Farquhar et al. (1980)
Farquhar et al. (1980)
Farquhar et al. (1980)
Rogers and Aneja (1980)
van Hove et al. (1987)
van Hove et al. (1987)
van Hove et al. (1987)
van Hove et al. (1987)
Hutchinson et al. (1972)
Rogers and Aneja (1980)
Rogers and Aneja (1980)
Duyzer et al. (1987)
Lockyer and Whitehead
(1986)
Lockyer and Whitehead
(1986)
Hutchinson et al. (1.972)
Rogers and Aneja (1980)
Rogers and Aneja (1980)
Rogers and Aneja (1980)
van Hove et al. (1989)
van Hove et al. (1989)
Hutchinson et al. (1972)
Rogers and Aneja (1980)
Hutchinson et al. (1972)
Rogers and Aneja (1980)
Rogers and Aneja (1980)
Rogers and Aneja (1980)
Tor NH3 at 25 °C 1 Mg/rn3 = 0.00143 ppmv.
bData are the mean or a range of reported values.
°K! is based on a one-sided leaf area.
dNA = not available.
August 1991
10-27
DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-9. DEPOSITION VELOCITY (Vd) OF NH3
TO PLANT CANOPY SURFACES
(Data showing net efflux of ammonia from fertilized crop landscapes are not included in this table.)
Species
Bean (Phaseolus vulgaris)
Fescue
(Festtica arundinaced)
Heather/purple moor grass
(Calluna/Molina)
Maize (Zea mays)
Oats (Avena sativd)
Orchard grass
(Dactylis ghmerata)
Pine (Pimis sp.)
Soybean
(Glyclne max [L.] Merr.)
Concentration
ppmv (/ig/m-3)a
0.100
, 0.603
NAd
0.250
0.200
'0.576
NA
0.075
[VJ"0
mm s"1
4
12
19
3
10
10
18-26
6
Method
Chamber
Chamber
Flux grad.
Chamber
Chamber
Chamber
Flux grad.
Chamber
Reference
Aneja et al. (1986)
Aneja et al. (1986)
Duyzer et al. (1987)
Aneja et al. (1986)
Aneja et al. (1986)
Aneja et al. (1986)
Duyzer et al. (1987)
Aneja et al. (1986)
'For NH3 at 25 °C 1 jig/m'3 = 0.00143 ppmv.
bData are means or a range of reported values.
"Data are based on ground area under the canopy.
dNA = not available.
1
2
3
4
5
6
7
8
9
10
11
12
13
demonstrate cuticular deposition rates (see the discussion in Section 9.4.1). However,
cuticular deposition rates are 1 to 2 orders of magnitude less than representative stomatal
uptake rates for tree foliage. Because cuticle deposition is low it should be considered of
minor importance, but not ignored when calculations of total N deposition to landscapes are
attempted.
Whole-canopy measurements of NO2 deposition conducted in laboratory or field
situations (Table 10-3) yield daytime overall deposition velocities (Vd) between 1 and
28 mm s'1. Duyzer et al. (1983) and van Aalst and Diederen (1985) cautioned that field
measurements of NO2 deposition: may have been in error because NO2 analyzers are also
sensitive to HNO3 vapor. Nitric acid vapor has a higher deposition velocity than NO2
(Section 10.2.4.3) and if monitored simultaneously with NO2 could have resulted in an
overestimate of deposition (e.g., tffill, 1971). Chemical reactions resulting from
photochemical reactions between NO, NO2, and 03 can also lead to errors in whole-canopy
August 1991
10-28 DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-10. MEASURED DEPOSITION
VELOCITIES OF NITRATE (NO3') AND AMMONIUM (NH4+)
Deposition Velocity
Species
American elm
N03"
11
NH4+
NA
Method
Leaf wash
Reference
Dasch (1987)
(Ulmus americana)
Austrian pine
(Pinus nigra)
Beech (Fagus silvatica)
winter
Ceanothus crassifolius
5-13
13
7-17
6-16
4.T
0.1-0.6 Leaf wash Dasch (1987)
10 Throughfall Hoefken and Gravenhorst (1982)
6-13 Throughfall Gravenhorst et al. (1983)
2-8 Throughfall Gravenhorst et al. (1983)
4.4 Leaf wash Bytnerowicz et al. (1987)
Chestnut oak (Quercus prinus)
dormant
Heather/moor grass
(Calluna/Molina)
Laurel
(Kalmia latifolia)
Norway spruce
(Picea abies)
5.5
7.1
NAb
NA
NA Throughfall Lovett and Lindberg (1984)
NA Throughfall Lovett and Lindberg (1984)
1.8 Flux grad. Duyzer et al. (1987)
0.3-1.4 Leaf wash Tjepkema et al. (1981)
winter
Pasture land
Pin oak
(Quercus palustris)
Privet
(Ligustrum japonicwri)
(Ligustrum ovalifolium)
Soybean (Glycine max)
White pine
(Pinus strobus)
11-37
13-32
7-8
7-11
2.2-5.4
1-2.1
2.4
NA
7-21
6-16
NA
NA
NA
NA
NA
0.3-1.4
Throughfall
Throughfall
Gradient
Leaf wash
Leaf wash
Leaf wash
Leaf wash
Leaf wash
Gravenhorst et al. (1983)
Gravenhorst et al. (1983)
Huebertetal. (1988)
Dasch (1987)
John et al. (1985)
John etal. Q985)
Dolske (1988)
Tjepkema et al. (1981)
"Particle NO3" deposition data typically includes some NO3" from HNO3 vapor.
bNA = not available.
August 1991
10-29
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
Vd measurements based on micrometeorological techniques (Hicks et al., 1989). Delany
et al. (1986) reported that eddy correlation measurements conducted over a grassland were
not appropriate for measurements of the fluxes of nitrogen oxides. Their data showed that
deposition of NOX predominated in the morning hours while emissions of NOX were observed
in the afternoon. However, their results which include both NO and NO2 were confounded
by photochemical reactions with ozone, resulting in the bimodal pattern of diurnal deposition.
Hicks and Matt (1988) also measured apparent bidirectional fluxes of NO2 from forest
canopies, but they could not conclude that such fluxes were a consequence of natural NO2
emissions (i.e., anthropogenic sources of NO2 and/or in-canopy transformations of NO2 to •
NO could have been responsible for the observed data). Fitzjarrald and Lenschow (1983)
conclude that the deposition velocity (Vd) concept is invalid for circumstances when chemical
reaction time is less than or comparable to the time required for turbulent diffusion. It
appears that this may often be the case for micormeteorologically based measurements of
canopy NO2 deposition.
The leaf-level measurements of NO2 deposition presented in Table 10-2 encompass a
large number and type of plant species. A simple average of the species specific data in
Table 10-2 for non-dormant plants indicates the following trend for deposition of NO2:
broadleaf trees = crop plants > conifer trees = house plants. Mean leaf conductance to
NO2 (Kj) for broadleaf tree and crop plants was approximately 1.3 mm s"1, and for conifers
and house plants between 0.5 and 1.0 mm s"1. Hanson et al. (1989) documented a similar
pattern. ElMey et al. (1982) reported data on the foliar sorption of NO2 to ten ornamental
woody plants using an NO2 concentration of 400 nl I"1. Based on one-sided leaf areas for
conifers, they observed higher NO2 deposition to conifers than to hardwoods. Had they used
total area to normalize their conifer data it would have shown the opposite pattern. Okano
et al. (1988) reported a positive correlation between NO2 deposition and stomatal conductance
for eight different crops which followed a trend associated with stomatal densities of the
foliage. Grennfelt et al. (1983) also found a strong relationship between NO2 deposition and
stomatal conductance for Scots pine.
August 1991,
10-30
DRAFT-DO NOT QUOTE OR CITE
-------
1 10.2.4^2 Nitric Oxide
2 • A comparison of tree and crop data between Tables 10-2 and 10-4, or Tables 10-3 and
3 10-5 shows that the Kx and Vd of NO are considerably less than for NO2. Lower
4 - conductance and deposition velocities indicate a reduced potential for the deposition of NO by
5 ! leaves than for NO2. The lower rate of deposition for NO is expected because of NO's
6 - lower aqueous solubility. Deposition data for several species of "house plants" reported by
7 Saxe (1986) indicated the same trend. The deposition of NO to foliar surfaces increased in a
8 linear manner with respect to ambient concentrations (Skarby et al., 1981), and stomatal
9 control over NO deposition has been documented by Saxe (1986). Kisser-Priesack et al.
10 (1987) also documented the capacity of Norway spruce and tomato cuticles to absorb gaseous
11 ' NO labeled with 15N, and concluded that a cuticular pathway for foliar deposition should not
12 be ignored.
13 , As for NO2, a compensation point for NO deposition to leaves has been indicated.
14 Nitric oxide concentrations greater than 0.05 ppmv routinely lead to deposition onto plant
15 canopies (Tables 10-4 and 10-5), but NO has also been observed to be evolved from foliage
16 (Farquhar et al., 1983). Klepper (1979) measured NO evolution from soybean plants stressed
17 with herbicides, and an enzyme system responsible for the conversion of nitrite to nitrogen
18 oxides has been described by Dean and Harper (1988). Nitric oxide emissions from plants
19- ••. 'are not widespread; and have only been documented completely for a specific set of plants in
20 the bean family (Leguminosae) (Dean and Harper, 1986).
21 Although more research is needed, two alfalfa studies suggest low deposition velocities
22 for NO to plant canopies (Table 10-5). Given NO's potentially greater phytotoxicity (see
23 - Section 9.4.3) deposition data from a broader array of plant species is needed.
24 •:..•-•
25 10.2.4.3 Nitric Acid Vapor <
26 The dry deposition characteristics of HNO3 vapor suggest substantially higher deposition
27 for HNO3 than for other nitrogen oxides. Micrometeorological measurement of the overall
28 deposition velocity of HNO3 to pasture grass (see papers by various authors in Table 10-7),
29 showed an average Vd for HNO3 of 29 mm s"1. Other studies on crop canopies showed Vd
30 values for HNO3 over a range from 4-260 mm s'1. Using throughfall nitrate and ambient
31 HNO3 concentrations, Dasch (1987) calculated the Vd for an Austrian pine (Pinus nigra)
August 1991
10-31 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
stand to be 67 mm s"1 at the stand perimeter and 17 mm s"1 at interior stand locations.
Dollard et al. (1987) reported Vd values as high as 260 mm s"1 for wheat canopies, but
recent modeling efforts (Bennett, 1988; Meyers and Hicks, 1988; Meyers et al., 1989)
indicate that such high Vd levels may not be possible. Fowler et al. (1989a) assumed HNO,
•?
and HC1 deposition to vegetation landscapes to be similar and concluded that Vd values for
low stature vegetation and crops would range from 5 to 50 mm s"1 depending on wind speeds
(Table 10-7). Forest landscapes also showed a range of Vd from 40 to 100 mm s"1 for low
and high wind speeds, respectively.
A computer model and ambient HNO3 concentrations were employed by Hicks et al.
(1985) to predict the Vd of HNO3 to broadleaf and high elevation red spruce forests. Their
analysis predicted a Vd between 20 and 50 mm s"1 for the low elevation broadleaf forests,
and a Vd between 60-120 mm s"1 for red spruce forests at high elevations. However, a more
recent simulation for crop canopies (Meyers and Hicks, 1988) projected that HNO3
deposition rates are mainly limited by the atmosphere-canopy turbulent exchange mechanisms
(wind), and predicted Vd values between 5 and 20 mm s"1 for slow and fast wind speeds,
respectively. Fowler (1984) calculated that the atmospheric resistance to deposition of
pollutants would increase from 2 to 4 fold depending on the nature of the landscape
vegetation with a change in windspeed from 1 to 4 m s"1. Flux gradient simulations based on
weekly mean filter pack HNO3 concentration measurements for'a deciduous forest canopy
(Meyers et al., 1989) showed 35 mm s"1 to be an appropriate mean Vd with a range between
20 and 60 mm s"1.
Only a few studies have attempted to measure HNO3 deposition to individual leaves.
Dasch (1989) used a mass balance approach to measure HNO3 deposition to tree foliage
(Table 10-6) and found a mean ;KX for two hardwoods to be 8.2 mm s"1 and a Kt for
Pinus nigra to be 2 mm s"1. Marshall and Cadle (1989) also used a mass balance approach
to measure HNO3 dry deposition to dormant pine shoots and found much lower Kx values
ranging from 0.4 to 0.8 mm s'1. Hanson et al. (1991) measured HNO3 conductances to
foliage of four tree species under low humidity conditions and found a Kj ranging from
1 to 3.3 mm s"1. Because low humidity caused stomatal closure, their measurements did not
include deposition to leaf internal spaces. Vose and Swank (1990) used a 15N labeling
technique to assess HNO3 deposition to white pine foliage and found rates of
August 1991
10-32 DRAFT-DO NOT QUOTE OR CITE
-------
1 "non-extractable" HNO3 absorption between 5 and 53 limol g"1 s"1. These data were not
2 included in Table 10-6 because the surface adsorbed HNO3 was removed in a water rinse
3 prior to assaying nonextractable 15N-labeled HNO3. Taylor et al. (1988) compared foliar
4 deposition characteristics of HNO3 vapor to that of other pollutant gases and suggested that
5 HNO3 deposition might be predominantly to the cuticle. This contrasts with patterns for NO
6 and NO2 which show most deposition to leaf interiors. Hanson and Taylor (1990) modeled
7 dry deposition of four pollutant gases to a hypothetical leaf surface, and predicted that HNO3
8 vapor deposition through plant cuticles would be greater than cuticular deposition of NO, O3,
9 and SO2. Vbse and Swank (1990) conducted a study of HNO3 deposition to foliar surfaces
10 using 15N labeled HNO3 has confirmed the cuticular pathway for HNO3 deposition.
11
12 10.2.4*4 Ammonia
13 Ammonia deposition data is limited primarily to crop plants. The average deposition
14 variables for all crop species included in Tables 10-8 and 10-9 are a Kj for leaves of
15 5.6 mm s"1 and a Vd for canopies of 7.4 mm s"1. Rates of NH3 deposition at concentrations
16 above 0.01 ppmv are linearly related to ambient concentration levels (van Hove et al., 1987;
17 Porter et al., 1972). However, Farquhar et al. (1980) observed a temperature dependent
18 evolution of NH3 from bean plants resulting in no net exchange of NH3 at ambient
19 concentrations between 0.003 and 0.005 ppmv. For ambient concentrations below that
20 "compensation point" NH3 evolution was observed, and above that concentration NH3 was
21 deposited in proportion to ambient NH3 concentrations. Lemon and van Houtte (1980) used
22 micrometerological techniques to reach similar conclusions (i.e., net NH3 deposition is
* "
23 concentration dependent).
24 Limited data for forest species show a similar range of Kx and Vd values. Duyzer et al.
25 (1987) has reported Vd for ammonia to heather-purple moor grass (Calluna, Molinia sp.)
26 canopies to be 19 mm s"1, and Vd to Corsican pine (Pinus nigra var. maritime) canopies
27 ranged between 18 and 26 mm s"1. These values are somewhat greater than those predicted
28 for crop plants. Van Hove et al. (1989a) found that NH3 deposition to Phaseolus vulgaris
29 and Populus euramericana cuticles decreased with decreasing relative humidity. Furthermore,
30 the cuticle deposition sites exhibited saturation given sufficient exposure time; little of the
31 adsorbed NH3 appeared to pass through the cuticle. However, cuticular deposition of NH3
. August 1991
10-33
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
represents only about 3% of the amount taken up via the stomata (van Hove et al., 1989a).
Van Hove et al. (1989b) reported additional Kj data for internal and external surfaces of
P. euramericana leaves ranging from 0.5 to 9 mm s"1 depending on stomatal conductance.
Van Hove et al. (1990) concluded that calculation of NH3 deposition to leaves using only
stomatal conductance data could result in a serious underestimation of the flux for conditions
of low temperature and high relative humidity.
Diurnal patterns of NH3 deposition follow similar patterns as for plant CO2 uptake
(Hutchinson et al., 1972). Other studies have related NH3 deposition to diurnal patterns of
stomatal opening (Aneja et al.,; 1986; Rogers and Aneja, 1980). A net deposition of 21 and
86 #mol g fresh weight"1 h"1 at 30 and 300 ppmv, respectively, was measured in sunflower
leaves using high concentrations of 15N labeled NH3 (Berger et 4.,, 1986).,: 15N labeled NH3
was incorporated into corn seedlings (Porter et al., 1972). Numerous other papers
encompassing a range of plant species indicate that NH3 exchange between crop canopies and
the atmosphere is a dynamic process, and concentration gradients between the atmosphere and
the landscape determine whether net influx or efflux of NH3 will take place (alfalfa—Dabney
and Bouldin, 1985; grazed pasture—Denmead et al., 1974; maize—Farquhar et al.-, 1979;
wheat—Harper et al., 1983, 1987; Parton et al., 1988). All of these studies involved ,some
type of fertilization regime, and it remains unclear to what extent "nutrient poor" .natural
ecosystems might exhibit NH3 efflux.
Modeling simulations have come to similar conclusions.. A modeled "canopy-level" Vd
for ryegrass (Lolium perenne L.) was reported to be 3 to 14 mm s"1 (Cowling and Lockyer,
1981). Sinclair and van Houtte (1982) simulated the deposition of NH3 to a soybean canopy
and determined that significant foliar deposition would occur at ambient concentrations as low
o
as 1 #g m . However, net deposition of NH3 by the combined soil-vegetation landscape was
predicted to occur routinely only at NH3 concentrations in the range from 40-70 fig m"3.
Denmead et al. (1976) found that ungrazed pasture was capable of absorbing most NH3
released from the ground while :grazed pasture lost NH3 to the atmosphere. Their
observations, while not quantitative, suggest that foliage of an ungrazed grass-clover pasture
is an effective sink for soil generated NH3. Denmead et al. (1978) demonstrated that a corn
field (Zea mays) exhibited net absorption of NH3 only when soil surfaces were dry.
August 1991
10-34
DRAFT-DO NOT QUOTE OR CITE
-------
1 10.2.4.5 Particles (Nitrate and Ammonium)
2 Direct measurements of aerosol-associated N deposition to foliar and inert surfaces have
3 been based on surface extractions, and extrapolations of throughfall information.
4 Unfortunately, these types of observations are of limited value due to the inability to separate
5 aerosol NO3" and NH4+ deposition from deposition due to HNO3, NO2, and NH3 that
6 display the same ionic forms once deposited to landscape 'surfaces (Bytnerowicsz et al., 1987;
7 Dasch, 1987; Lindberg and Lovett, 1985; van Aalst and Diederen, 1985). The average Vd
8 for nitrate and ammonium (Table 10-10) was greater if determined from throughfall '
9 measurements (12 and 10 mm s"1) than if determined from individual leaf washing
10 experiments (6 and 2 mm s"1). However, these differences in Vd between measurement
11 techniques are primarily a function of scale. The leaf wash measurements extract adsorbed
12 ions from a defined leaf area, but throughfall measurements extract ions from all layers of the
13 canopy (an undefined area) and relate it only to the ground area of the stand (see also
14 discussion in Section 10.2). Lindberg and Lovett (1985) estimated dry deposition of nitrate
15 to deciduous forest leaves to be 5.7 /xg ni"2 h"1, but declined to calculate a deposition velocity
16 because of difficulties in (1) obtaining accurate particulate NO3~ air concentrations (Appel and
17 Tokiwa, 1981) and (2) the contribution of HNO3 dry deposition to NO3" on the foliage
18 surface could not be separated from that of aerosol NO3". Dolske (1988) reported Vd values
19 for NO3" deposition to soybean to range from 30.8 down to 0.4 with a mean of 2.4 mm s" .
20 However, because Dolske's leaf wash measurements included a component of HNO3 vapor
21 "'• the Vd values may represent more than deposition due to' aerosol nitrate alone. •
22 Only one published paper has used micrometeorological methods to determine the
23 aerosol nitrate and ammonium deposition to landscape surfaces. The Vd information from
24 Duyzer et al. (1987) for aerosol NH4+ deposition to heathlands (1.8 mm s"1, Table 10-10)
25 was determined using flux gradient analysis' of NH4+ particles trapped in filtered air leaving
26 denuder tubes.
27 • '
28 10.2.5 Deposition of "N" Forms to Non-Foliar Surfaces
29 In addition to foliage, deposition of particles and gases has also been measured to bark,
30 soil, and snow covered surfaces (Table 10-11). Measured NO2 deposition of NO2 to normal
31 or wetted bark of three broadleaf and one conifer tree species was similar among species
August 1991
10-35
DRAFT-DO NOT QUOTE OR CITE
-------
SURFACES TO REACTIVE NITROGEN GASES
Species
Nitrogen dioxide
Forest floor
Hardwood
Conifer
Bark
dry
wet
Forest litter
Hardwood
Conifer
Soil
Waltham, MA
sandy loam
adobe clay
Oak Ridge,TN
forest
Snow
Nitric Oxide
Soil
sandy loam
adobe clay
forest soil
Snow
Nitric Acid Vapor
Bark
Snow
-18 °C
-8
-5
-4
-3
-2
Concentration
ppmv Gitg/m"3)
0.044
0.043
0.066
0.058
0.076
0.074
3-100
13-53
13-53
0.050
NA
0.006-0.03
11-4
1-4
NA
0.0005-0.002
0.06-0.07
0.014 (36)a
0.014 (36)
0.014 (36)
0.014 (36)
0.014 (36)
0.014 (36)
[Ki]ab
mm s
4.7
4.8
0.47
0.93
0.06
-0.05
0.2
6
7.7
' 4.2
3.0
«0.3
1.9
1.3
<0.01
«0.3
7.4
<0.2
0.4
0.4
1.2
1.0
5.7
Method
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
NA
Chamber
Chamber
Chamber
NA
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Chamber
Reference
Hanson et al. (1989)
Hanson et al. (1989)
Hanson et al. (1989)
Hanson et al. (1989)
Hanson et al. (1989)
Hanson et al. (1989)
Abeles et al. (1971)
Judeikis and Wren (1978)
Judeikis and Wren (1978)
Hanson et al. (1989)
van. Aalst (1982)
Granat and Johansson (1983)
Judeikis and Wren (1978)
Judeikis and Wren (1978)
van Aalst (1982)
Granat and Johansson (1983)
Hanson et al. (1991)
Johansson and Granat (1986)
Johansson and Granat (1986)
Johansson and Granat (1986)
Johansson and Granat (1986)
Johansson and Granat (1986)
Johansson and Granat (1986)
'Data are presented as a range or the mean of reported values.
bData are based on total area for bark and litter, and ground area for snow, forest floor, and soil.
August 1991
10-36
DRAFT-DO NOT QUOTE OR CITE
-------
1 (Hanson et al., 1989). The conductance of NO2 to wet bark was almost double that to dry
2 bark (Table 10-11). The conductances (Kt) ranged from 0.44 to 0.84 mm s"1 and were
3 within a factor of 2 of Kj values for plant leaf surfaces. HNO3 conductance to bark was
4 nearly an order of magnitude greater than for NO2 (Hanson et al., 1991; Table 10-11). No
5 data are available for the deposition of other forms of dry deposited nitrogen to bark.
6 The deposition velocity of NO2 to soil exceeds that for NO (Table 10-11;'JudeiMs and
7 Wren, 1978). When compared to foliage or bark surfaces, deposition to the forest floor and
8 soil surfaces show a disproportionately high rate (compare data from Tables 10-2 and 10-11).
9 A comparison of deposition to the soil and forest showed that soil was the primary receptor
10 site of NO2 (Hanson et al., 1989). Abeles et al. (1971) measured NO2 deposition to fresh
11 and autoclaved soil and determined that a biological sink was responsible for approximately
12 12% of the soil NO2 deposition. However, Ghiorse and Alexander (1976) found no
13 difference in soil deposition after autoclaving or T-irradiation and concluded that
14 microorganisms were responsible, not so much for absorption of NO2, but its conversion into
15 nitrate. Mortland (1965) and Sundaresan et al. (1967) documented mechanisms for NO
16 deposition by soil based on adsorption or interaction with soil minerals. Prather et al. (1973)
17 and Prather and Miyamoto (1974) provided data on the deposition of NO2 and NO to
18 calcareous soils, but these data are not included in Table 10-11 because of the extremely high
19 air concentrations used (0.1 to 1.5% by volume).
20 Nitric acid vapor is the only oxide of nitrogen to exhibit significant deposition to snow,
21 but it does so only when temperatures exceed -5 °C (Table 10-11; Granat and Johansson,
22 1983; Johansson and Granat, 1986). Bennett (1988) modeled the deposition of reactive
23 gases, such as HNO3, to urban environments (i.e., city scapes) and calculated that Vd would
24 be limited to 2-5 mm s"1 by aerodynamic resistances.
25
26
27 10.3 EFFECTS ON VEGETATION AND SOILS
28 10.3.1 Introduction
29 . The effects of any nutrient upon biological systems must be viewed from the perspective
30 of the amount of that nutrient in the system, the biological demand for that nutrient, and the
31 amount of input. Thus, if a nutrient is deposited on an ecosystem deficient in that nutrient, a
August 1991
10-37
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
growth increase will occur, and this will be generally (but not always) be regarded as a
positive effect (the deficiency condition in Figure 10-3). If a nutrient is deposited on an
ecosystem with adequate supplies of that nutrient, there may be no effect for a period of time
or over a range of input values (the sufficiency condition in Figure 10-3). Inputs of any
nutrient greatly in excess of an ecosystem's biological demand will result in negative growth
responses, or toxic effects of some sort, as shown in the last segment of the curve in
Figure 10-3.
Nitrogen (N) is unique among nutrients in that its retention and loss is regulated almost
exclusively by biological processes. Whereas other major nutrients (P, S, K, Ca, Mg, Mn)
originate primarily from soil minerals and often accumulate in adsorbed/exchangeable pools
in the soil, nitrogen originates from the atmosphere and rarely accumulates for long in
exchangeable/adsorbed pools. (Ammonium may accumulate by fixation in the interlayers of
2:1 clays or by chemical reactions with humus, but these pools are largely unavailable to
either plants or microbes.) In theory, large soil pools of NH4+ could occur, because NH4+
strongly adsorbs to cation exchange sites (negatively-charged sites on clays and organic matter
in soils). Large soil NH4+ pools seldom occur, however, because of the action of nitrifiers
(soil organisms that convert NH4+ to NO3", a process referred to as nitrification), and, in
alkaline soils, purely chemical conversion to NH3 gas followed by volatilization. In those
rare soils where nitrification is inhibited and acidity is too great for volatilization, soil NH4+
pools can build up to fairly high levels (e.g., Roelofs et al., 1987; Vitousek et al., 1979), but
these cases seem to be the exception rather than the rule. Since NO3" is poorly adsorbed to
oils (in contrast to SO42" and ortho-phosphate; Kingston et al., 1967), nitrification in excess
of plant and microbial demand for N almost always leads to increased NO3" leaching (e.g.,
van Breemen et al., 1982; Van Miegroet and Cole, 1984; Johnson and Todd, 1988; Foster
and Nicolson, 1988). High rates of NO3" leaching can be deleterious for two major reasons:
(1) the potential acidification of soils and waters and/or mobilization of A13+ (as is the case
i
with SO42"; Reuss and Johnson, 1986) and (2) the potential contamination of drinking water
(the EPA standard for NO3" -N being 10 mg N/L).
Soils are by far the largest N pool in forest ecosystems, usually exceeding 85% of total *
ecosystem capital (Cole and Rapp, 1981). Yet most soil N is inert and unavailable for either ;
uptake or leaching, with only a rather loosely-defined "mineralizeable" pool being
August 1991
10-38
DRAFT-DO NOT QUOTE OR CITE
-------
o
DC
Q.
DEFICIENCY
SUFFICIENCY
TOXIC
NUTRIENT SUPPLY
Figure 10-3. Schematic representation of the response of natural ecosystems to nutrient
inputs.
1 biologically active (Aber et al., 1989). This "mineralizeable pool", the size of which is
2 -., typically defined either by in situ incubation of soils or litter, is that portion of soil N that
3 heterotrophs (decomposers), autrotrophs (plants), and nitrifying bacteria compete for. The
4 processes involved in this competition have been described and modeled, often with a special
5 emphasis on nitrification and nitrate leaching (e.g., Vitousek et al., 1979; Riha et al., 1986).
6 , However ^ a generally applicable and potentially predictive model analogous to, for example,
7 cation exchange and leaching (e.g., Reuss, 1983; Gherini et al., 1985; Cosby et al., 1985)
8 remains elusive. For example, the cessation of nitrate leaching following harvesting in
9 N-rich red alder (Alnus rubrd) forests in Washington (apparently a result of cessation of
10 N-fixation; Biggar and Cole, 1983; Van Miegroet et al., 1989) does not support earlier
11 predictions that nitrate leaching following disturbance is usually greatest in sites with
12 inherently better N status (e.g., Vitousek et al., 1979). Also, the recent discovery of several
13 sites where nitrate leaching is high under undisturbed conditions (Van Miegroet and Cole,
14 1984; Foster, 1985; Joslin et al., 1987; Johnson et al., in press) does not support the long-
15 held notionfthat nitrogen is tightly cycled and conserved in forest ecosystems (e.g., Gessel
16 et al., 1973; Cole and Rapp, 1981; Aber et al., 1989).
August 1991
10-39
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
The following discussion will focus upon forest ecosystems, because of their sensitivity
to both the positive and negative effects of N deposition discussed above. Arid and semi-arid
ecosystems are not as susceptible to the soil acidification and groundwater NO3~ pollution as
are forest and agricultural systems in more humid areas because of a lack of water for NO3"
leaching and because soils are more alkaline. There are some important implications of
N deposition on arid and semi-arid ecosystems, however, that deserve consideration; namely,
vegetation growth increases and increased denitrification. Therefore, due consideration of
nitrogen cycling in and nitrogen deposition effects on arid ecosystems is given where
information is available.
Agricultural lands are excluded from this discussion because crops are routinely
fertilized with amounts of N (100-300 kg/ha) that far exceed pollutant inputs even in the most
heavily polluted areas. These high rates of fertilization can lead to groundwater
contamination problems and may contribute to the atmospheric N2O loading as well (e.g.,
Hutchinson and Mosier, 1979), but a discussion of the environmental effects of fertilization
are beyond the scope of this Section.
10.3.2 Pollutant N Inputs and Nitrogen Cycling in Natural Ecosystems:
A Brief Review
An evaluation of the effects of pollutant N deposition on terrestrial vegetation and soils
must begin with considerations of how these pollutant inputs affect terrestrial N cycles. The
general subject of terrestrial N cycling was reviewed in section 10.1.3; only a few of the
more germane details are repeated here. Not all potential effects of pollutant N inputs can be
described from the perspective of a generic N cycle, however: there are secondary effects
such as effects of excessive vegetation uptake upon susceptibility to attack by pests and
pathogens and to climatic damage, for example, that cannot be described from the N cycle
alone.
Nitrogen, unlike Ca, K, Mg, P, or S, seldom forms large soil inorganic pools which
can buffer excessive inputs and provide a readily-available source of nutrient for plants. In
theory, large soil pools of NH4+ could occur, because NH4+ strongly adsorbs to cation
exchange sites. Large soil NH4+ pools seldom occur, however, because of the action of
nitrifiers, and, in alkaline soils, purely chemical conversion to NH3 gas followed by
August 1991
10-40
DRAFT-DO NOT QUOTE OR CITE
-------
1 volatilization. In those rare soils where nitrification is inhibited and pH is too low for
2 volatilization, soil NH4+ pools could, in theory, build up to fairly high levels (e.g., Roelofs
3 et al., 1987; Vitousek et al., 1979), but these cases seem to be the exception rather than the
4 rule. The potential for the accumulation of large NH4+ pools can also be reduced by purely
5 chemical reactions between ammonium and soil humus (e.g., Foster et al., 1985). Since
6 NO3" is poorly adsorbed to soils, nitrification in excess of plant and microbial demand for N
7 almost always leads to increased NO3" leaching (e.g., van Breemen et al., 1982; van
8 Miegroet and Cole, 1984; Johnson and Todd, 1988; Foster and Nicolson, 1988).
9 Nitrogen can enter forest ecosystems in many forms: (1) wet deposition of NH4+,
10 NO3", and organic N; (2) dry deposition of these forms plus nitric acid vapor (Lindberg
11 et al., 1986), and (3) biological fixation of N2. Inputs via wet and dry deposition first
12 encounter the forest canopy where they may be taken up either by trees or by organisms
13 living within the canopy, or the phyllosphere (leaf surface). Deposited N not taken up within
14 the phyllosphere falls primarily as wet deposition to the forest floor, where plants,
15 decomposers (heterotrophs, which consist of fungi and bacteria), and nitrifiying bacteria
16 compete for it (Figure 10-4, top). This competition for N among heterotrophs, plants, and
17 nitrifiying bacteria plays a major role in determining the degree to which a vegetation growth
18 increase will occur and the degree to which incoming N is retained within the ecosystem. It
19 has been assumed that nitrifiers are poor competitors for N compared to heterotrophs and
20 plants (Vitousek et al., 1982; Riha et al., 1986; see also review by Davidson et al., 1991).
21 This assumption has recently been challenged by Davidson et al. (1991). Using 15N
22 techniques, these authors found significant nitrification and microbial NO3" uptake (12 to
23 46% of N mineralization rates) in grassland soils even when soil NO3" pools and NO3"
24 leaching rates were very low. They concluded that the small soil NO3" pool in this site
25 turned over very rapidly due to nitrification and microbial uptake NO3" and that nitrifiers
26 were quite able competitors for N. They also point out that NO3" production during
27 incubation actually represents a net effect of nitrification and microbial NO3" uptake, and
28 define this as "net nitrification". The extent to which these results might apply to forest
29 ecosystems is unknown; however, if this pattern proves to be true in general, it will require a
30 substantial redesign of the conceptual model currently used to explain and predict nitrification
31 and NO3" leaching.
August 1991
10-41
DRAFT-DO NOT QUOTE OR CITE
-------
LOW N DEPOSITION
VOLATILIZATION
[toy Plant*)
yl^j— I ";
I MlfUFCATIOM* ------
itrih«>) -r \
^^JT ^^.
"^ IMMCMILIZ*
IMMOBILIZATION
(by H»t»«oaofih[
uptoh* «nd chvfni
NH4 fxction with
FERTILIZATION
HIGH N DEPOSITION
Figure 10-4. Schematic representation of the fate of incoming N in N-poor (top),
fertilized (center), and high-N input systems.
August 1991
10-42 DRAFT-DO NOT QUOTE OR CITE
-------
1 Heterotroph demand for N (both NH4+ and NO3") depends upon the supply of labile
2 organic C substrates (as well as temperature and moisture conditions). Thus, adding labile
3 organic C to a soil should reduce plant uptake and net nitrification by increasing heterotrophic
4 competition for NH4+ and increasing microbial NO3" uptake. Adding labile organic C to a
5 soil may also cause increased activity of denitrifying organisms, which also require organic
6 substrates, resulting in reduced nitrate leaching. Turner (1977) demonstrated that addition of
7 carbohydrates to a forest soil in Washington caused increased N deficiency in Douglas fir
8 (Pseudotsuga menziesit) trees, presumably by stimulating heterotrophic competition for N.
9 Johnson and Edwards (1979) found that addition of carbohydrate substrate to a forest soil
10 caused an immediate reduction in nitrate leaching and net nitrification production during
11 laboratory incubation of a yellow-poplar forest soil in Tennessee.
12 According to the conceptual model described above, nitrification and NO3" leaching will
13 become significant only after heterotroph and plant demand for N are substantially satisfied, a
14 condition that has been referred to as "nitrogen-saturated". There are various definitions for
15 "nitrogen-saturation", many of which are reviewed by Skeffington and Wilson (1988). One
16 definition is "ecosystems where the primary production will not be further increased by an
17 increased in the supply of N". There are clearly problems with this definition in that
18 ecosystems that are low in N but limited by another nutrient (such as P) may not experience
19 an increase in primary production in response to N input unless P is added first (e.g.,
20 Pritchett and Comerford, 1982). Other definitions for nitrogen saturation reviewed by
21 Skeffington and Wilson (1988) include: "when external N input and N mineralization from
22 the soil exceed the capacity of the ecosystem organisms to absorb more N", or "an ecosystem
23 which cannot accumulate more N". Aber et al. (1989) define nitrogen saturation "as the
24 availability of ammonium and nitrate in excess of total combined plant and microbial
25 nutritional demand" (p. 379). This definition conveys the same idea as those reviewed by
26 Skeffington and Wilson (1988), but, in its strictest sense, it also is flawed. All ecosystems,
27 even extremely N-deficient ones, have some small pool of ammonium and nitrate within the
28 soil and litter components. Using the definition of Aber et al. (1989) in its strictest sense,
29 then, all ecosystems are nitrogen saturated to one degree or another. Aber et al. (1989) also
30 state that nitrogen saturation implies limitation on biotic function by some other resource
31 (e.g., phosphorus or water for plants or carbon for microbes). But if this is so, naturally
August 1991
10-43
DRAFT-DO NOT QUOTE OR CITE
-------
1 phosphorus (P)-deficient ecosystems (such as those in the southeastern coastal plain) might be
2 considered nitrogen saturated, whereas in reality these ecosystems are often very low in N
3 and release virtually no nitrate. Furthermore, as noted above, P-deficient ecosystems will
4 frequently accumulate substantially more N once P limitations are satisfied.
5 While the precise definition of nitrogen saturation seems elusive because of various
6 caveats that must be taken into account, the general idea seems to be encompassed in the last
7 and most brief definition reviewed by Skeffington and Wilson (1988): "an ecosystem which
8 cannot accumulate more N". This definition implies that further N accumulation cannot
9 occur, even though other nutrient limitations are satisfied. This definition will be used in the
10 following discussion.
11 It is important to note is that additional N inputs to an N-saturated ecosystem will cause
12 equivalent leaching losses of NO3" regardless of the chemical form of the N entering the
13 system (NH4+, NO3", or organic) to the extent that (1) N inputs are in biologically available
14 forms, (2) nitrification proceeds uninhibited, and (3) denitrification does not occur (Reuss and
15 Johnson, 1986). There has been an unfortunate tendency among atmospheric deposition
16 researchers to ignore the effects of NH4+ and (especially) organic N on ecosystem
17 acidification and nitrate leaching, an omission which substantially underestimates the
18 acidification potential of atmospheric N deposition.
19 The rather simple model depicted in Figure 10-4 does not account for the possibility of
20 nitrification inhibitors. Autotrophic nitrifiers are known to be inhibited by low pH high soil
21 solution Cl" concentrations, and certain organic chemicals, both naturally- and synthetically-
22 produced (Alexander, 1963; Roseberg et al., 1986). The occurrence and importance of
23 naturally-produced nitrification inhibitors has received considerable attention in the ecological
24 literature. An early study by Rice and Pancholy (1972) indicated that nitrification rates
25 decrease during forest succession due to the presence of chemical nitrification inhibitors
26 (soluble alleopathic compounds produced by plant litter). This somewhat controversial
27 finding stimulated several follow-up investigations in various ecosystems. Some of these
28 investigations supported the contention that nitrification inhibitors were a factor in controlling
29 NO3" losses from forest ecosystems (Lodhi, 1978; Olson and Reiners, 1983), but several
30 others found no evidence of them, and concluded that either competition for NH4+ or other
August 1991
10-44
DRAFT-DO NOT QUOTE OR CITE
-------
1 nutrient limitations controlled nitrification rates (Purchase, 1974; Robertson and Vitousek,
2 1981; Lamb, 1980; Cooper, 1986).
3 There is no reason to doubt that inhibitors play a role in some forests, but the extent to
4 which inhibitors occur and the factors leading to their production are unknown. Nor is it
5 known how inhibitors might function under conditions of very high, chronic NH4+ inputs.
6 Roelofs et al. (1987) report little nitrification in Dutch forests subject to very high inputs of
7 NH4+ from nearby agricultural activities, but they attribute the lack of nitrification in these
8 forests to low pH. The situation reported by Roelofs et al. (1987) is unusual, however; there
9 are few cases where these conditions do not lead to high rates of nitrification and NO3"
10 leaching. Others have reported high rates of nitrification under very acid soil conditions
11 (Klein etal., 1983; van Breemen et al., 1982, 1987).
12 Denitrification, (i.e., the microbially-mediated conversion of NO3" to NOX and N2
13 gases) is thought to be of importance only in forest soils which (1) have elevated NO3"
14 inputs, and (2) experience anaerobic conditions (e.g., flooded conditions) (Davidson and
15 Swank, 1987). Goodroad and Keeney (1984) provide estimates of denitrification losses from
16 relatively N-rich forest ecosystems in Wisconsin of 0.2 to 2.1 kg ha"1 yr"1, values that are
17 worthy of including in N budgets but do not compare to NO3" leaching rates that have been
18 shown to occur in some forests (see below). Similarly, Woodmansee (1978) discounts the
19 importance of denitrification in grassland soils, showing that NH3 volatilization from animal
20 wastes is the major N loss mechanism. Curiously, however, Westerman and Tucker (1978)
21 and Klubek et al. (1978) found that denitrification rather than NH3 volatilization is the major
22 N loss mechanism from desert soils in the Sonoran and Great Basin desert ecosystems. They
23 speculate that microsites with saturated water conditions occur during precipitation events that
24 produce the anaerobic conditions necessary for denitrification to occur. Peterjohn and
25 Schlesinger (1990) calculated that 77% of atmospheric N inputs to desert ecosystems in the
26 southwestern U.S. have been lost to the atmosphere since the last glaciation. They stop short
27 of giving values for NOX and N2 (denitrification) vs. NH3 (volatilization) losses, but point
28 out that the importance of learning more about the nature of gaseous N losses from these
29 systems, especially in the case of N2O given its importance to the ozone layer and as a
30 greenhouse gas.
August 1991
10-45
DRAFT-DO NOT QUOTE OR CITE
-------
1 Vegetation demand for N depends upon a number of growth-influencing factors
2 including temperature, moisture, and the availability of other nutrients. Limitation of
3 moisture in arid ecosystems clearly does not" preclude growth responses to N input, however.
4 Several studies have shown that demonstrated net N inputs to desert ecosystems produced
5 growth increases despite supposed water limitation? (Fisher et al., 1988c; see review by
6 Moorhead et al., 1986). Nitrogen is considered such an important factor in the productivity
7 and function of desert ecosystems that an entire volume has been devoted to the subject (West
8 andSkujins, 1978).
9 In Forest ecosystems, stand age is an important factor determining N uptake rates.
10 Uptake rates decline as forests mature, especially after the cessation of the buildup of
11 nutrient-rich foliar biomass following crown closure (Switzer and Nelson, 1972; Miller,
12 1981; Turner, 1981). Thus, one would expect NO3~ leaching rates to be greater in older
13 forests than in younger forests due to greater NH4+ supplies to nitrifiers as well as to lower
14 NO3" uptake in older forests. The results of Vitousek and Reiners (1975) support this
15 hypothesis in that they found higher NO3" concentrations in streams draining mature spruce-
16 fir forests than in streams draining immature spruce-fir forests in New England.
17 Processes that cause net N export from ecosystems such as fire and harvesting will
18 naturally push ecosystems toward a state of greater N demand or even N deficiency.
19 Frequent fire is normally thought of as an especially effective way of maintaining low
20 ecosystem N status. However, studies on the effects of fire upon soil N have produced
21 conflicting results. Some authors have reported total N contents that were not significantly
22 changed within 1-2 years of burning whereas others have reported significant losses.
23 Jurgensen et al. (1981) found that broadcast burning caused a minor net loss of
24 N (approximately 100 kg/ha) from a clearcut site in Montana, and concluded that plant
25 re-establishment benefitted from the increased N availability following this prescribed burn.
26 Wells (1971) noted that while the periodic prescribed burns has caused significant losses of
27 forest floor material immediately after the burn, there seemed to be a tendency for the system
28 to regain this organic matter over time and approach the control condition. He also found
29 that organic matter and nitrogen were redistributed from the forest floor to the surface
30 mineral soil as a result of burning, the net effect being a redistribution of the organic matter
31 in the profile rather than a reduction. Furthermore, one treatment (annually-burned plots)
August 1991
10-46
DRAFT-DO NOT QUOTE OR CITE
-------
\
1 showed significant increases in soil N (550-990 kg/ha), which were attributed to increased
2 activity of nitrogen-fixers. In contrast, Grier (1975) noted significant nitrogen losses
3 (855 kg/ha) from an intense fire on the eastern slope of the Cascade Mountains of
4 Washington. It seems that the net effect of fire on ecosystem N status has a great deal to do
5 with fire intensity.
6
7 10.3.3 Fate of Nitrogen in Forest Ecosystems: Contrasts Between Fertilizer
8 and Pollutants
9 The prospects for forests becoming "nitrogen saturated" from atmospheric N inputs have
10 been explored in recent workshops and reviews (Nilsson and Grennfelt, 1988; Schulze et al.,
11 1989; Aber et al., 1989). Critical loads analyses for N saturation typically consider
12 vegetation uptake and increment as the primary factors controlling forest ecosystem N
13 retention, and attribute little potential for soil N accumulation, despite the fact that soils
14 comprise the largest ecosystem N pool in virtually all forest ecosystems (Nilsson and
15 Grennfelt, 1988; Schulze et al., 1989). In contrast, numerous forest fertilization studies have
16 shown that litter and soils are major sinks for N (e.g., Heilman and Gessel, 1963; Mead and
17 Pritchett, 1975; Miller et al., 1976; Melin et al., 1983; Raison et al., 1990). As noted by
18 Aber et al. (1989), it is not surprising that forest ecosystems respond differently to pulse
19 inputs of N via fertilization vs. slow, steady inputs via atmospheric deposition.
20 Fertilization studies differ from pollutant N deposition in two important respects.
21 Pollutant N deposition enters the ecosystem at the canopy level whereas fertilizer is typically
22 (but not always) applied to the soil. Another important difference (as noted by Aber et al.,
23 1989) is that pollutant N deposition enters the ecosystem as a slow, steady input in rather low
24 concentrations, whereas the fertilizer is typically applied in 1-5 large doses. Nitrate
25 applications or urea applications to N-rich sites can result in substantial nitrate leaching losses
26 of fertilizer N (e.g., Overrein, 1969; Matzner et al., 1983; Tschaplinski et al., in press).
27 However, most studies show minimal loss of fertilizer N via leaching following single, large
28 applications of ammonium or urea to N-poor sites (Cole and Gessel, 1965; Overrein, 1969;
29 Cole et al., 1975; Worsnop and Will, 1980). As will be shown later, there are some
30 important differences in the way the nitrogen cycle in soils responds to large, single
31 applications vs. slow, steady applications of N, whether as fertilizer or as atmospheric input.
August 1991
10-47
DRAFT-DO NOT QUOTE OR CITE
-------
1 There have been cases where fertilizer has been applied in small, frequent doses, and it is
2 useful to briefly review some of those studies here before comparing fertilization with
3 atmospheric N deposition.
4 ,
5 10.3.3.1 Case Studies of Forest Fertilization at Differing Intervals
6 Ingestad (1980) has demonstrated in greenhouse experiments that optimum nitrogen
7 uptake and growth by plants can be achieved by adjusting N inputs to the rate of plant
8 growth. In these experiments the rate of N supply (i.e., flux density, or N input per unit
9 area per unit time) was proven to be the critical variable, not necessarily the concentration of
10 N in the uptake solution. Field experiments comparing standard fertilization with
11 simultaneous irrigation and fertilization (IF) have also demonstrated the superior growth
12 response and fertilizer N recovery by adjusting the flux density of N input (through the IF
13 treatments) as compared to adding either one or a few large doses of N as in conventional
14 fertilization (Aronsson and Elowson, 1980; Ingestad, 1980; Landsberg, 1986).
15 These authors (Aronsson and Elowson, 1980; Ingestad 1980; Landsberg, 1986) do not
16 report the effects of slow, steady inputs of N on nitrification and NO3" leaching. However,
17 multiple or continuous inputs of fertilizer may stimulate a buildup in populations of nitrifying
18 bacteria. A fertilizer experiment involving urea-N applications of 100 kg/ha/yr for 3 years in
. 19 quarterly (25 kg N/ha/3 mo) and annual (100 kg N/ha, in March) to young loblolly pine
20 (Pinus taedd) and yellow-poplar (Lirioderidron tulipiferd) plantations in very nitrogen-poor
21 sites in the Tennessee Valley (Johnson and Todd, 1988) found a buildup in nitrifying
22 bacteria. In all cases, the quarterly applications resulted in earlier and more pronounced
23 increases in soil solution nitrate than annual applications. Figure 10-5 illustrates this pattern
24 for the loblolly pine site. Furthermore, only the annual applications resulted in increased
25 growth (Figure 10-6, top). The authors concluded that more frequent fertilization in those
26 particular ecosystems benefitted nitrifiers more than trees.
27 In a later study, in a more N-rich site nearby, exactly the opposite results were obtained
28 in a study comparing a single urea-N applications of 50, 150, and 450 kg N/ha with multiple
29 (three times at 37.5 kg N/ha) applications to a young sycamore (Platanus occidentalis)
30 plantation (Tschaplinski et al., in press). In this case, the authors found much higher soil
31 solution NO3" concentrations in general (including in the control plots), no delay in the onset
August 1991
10-48
DRAFT-DO NOT QUOTE OR CITE
-------
QRNL-DWG 87-1479
20
15 —
10
T—r—m—rn—i i i t i
LOBLOLLY PINE
NON-MYCORRHIZAL
CONTROL
—*.•_*_* I
20
15
ca
I 10
in
o
Z 5
20
15
10
1 E I
I I I t I
ANNUALLY
«*•
*
...•.**,
i i i
QUARTERLY
_1_
JAN APR AUG NOV FEE MAY SEP DEC MAR JUL OCT JAN APR AUG NOV FE8
1982 | 1983 | 1984 | 1985 J1986
Figure 10-5. SoU solution nitrate concentrations in untreated control (top), annually
fertilized (100 kg urea-N ha"1 yr'1, center) and quarterly-fertilized (25 kg
urea-N ha"1 3 mo"1, bottom) loblolly pine plots. (After Johnson and Todd,
1988.) •"•«•-•
August 1991
10-49 DRAFT-DO NOT QUOTE OR CITE
-------
Stem Weight in Loblolly Fertilizer Study
•(000
3000-
2000-
1DOO-
Column 20
Stem Weight in Sycamore Fertilizer study
120
Figure 10-6. Top: Growth of loblolly pine in untreated (C), annual (A) (100 kg urea-N
ha"1 yr"1, center) and quarterly (Q) (25 kg urea-N ha"1 3 mo"1, center)
applications of urea-N. (After Johnson and Todd, 1988.) Bottom: Growth
of American sycamore in untreated (C), multiple (37.5 kg urea-N ha"1,
3 times) and single (450 kg N ha"1) applications of urea-N. (After
Tschaplinski et al., in press.)
August 1991
10-50
DRAFT-DO NOT QUOTE OR CITE
-------
1 of nitrate leaching, and the greatest rates of nitrate leaching in the single 450 kg/ha
2 application (Figure 10-7). Tree growth response was also greatest in the 450 kg/ha
3 treatment, but growth responses were also significant in the multiple fertilization treatment
4 (Figure 10-6, bottom). Thus, in this N-rich site, single fertilization produced the greatest
5 growth response, but at a higher cost in* terms of nitrate leaching.
6 The key to differences in nitrate leeching response observed in these two studies was the
7 initial relative abundance of nitrifiers. Aerobic incubations in the laboratory showed that the
8 delay period to the onset of nitrate production was 25-30 days in the N-poor site and 0-4 days
9 in the N-rich site (Johnson and Todd, 1988; Tschaplinski et al., in press). According to
10 Sabey et al. (1959) delay period for the onset nitrate production is closely related to the initial
11 population of nitrifying bacteria. These results imply that slow, steady inputs of N
12 characteristic of pollutant inputs may cause more rapid N-saturation in low-N ecosystems than
13 conventional, single-shot fertilization would, but the opposite would be true in high-N
14 ecosystems. If the initial population of nitrifiers is low, the 'slow, steady inputs will favor a
15 buildup of their populations more rapidly than single large inputs will and thus cause a
16 relatively early increase in nitrate leaching. If the initial population of nitrifiers is high, the
17 rate of nitrate leaching is more likely to be proportional to the input of N in excess of plant
18 demand regardless of timing and without delays caused by heterotrophic uptake.
19
20 10.3.3.2 Fate of N from Pulse Fertilization vs. Atmospheric Deposition
21 Forest fertilization has proven quite/successful in producing growth increases in
22 N-deficient forests, even though trees typically recover only 5-50% of fertilizer N in
23 aboveground biomass (the very high tree recovery found by Bockheim et al., 1986, being
24 exceptional; Table 10-12). Increased N in the soil is not mirrored directly by more N uptake
25 except at low levels (see Chapter 9). Fertilizer N retention in the litter and soil is usually
26 substantial (Table 10-12 and Figure 10-4, center). There are two possible mechanisms for
27 this high litter/soil N retention: (1) N uptake by soil heterotrophic organisms, and (2) non-
28 biological, chemical reactions between ammonia and soil organic matter (Foster et al.,
29 1985a). The overall result is that the retention of N on an ecosystem level is usually quite
30 high (averaging 60% of applied N; Table 10-12). Furthermore, fertilizer recovery in trees,
August 1991
10-51
DRAFT-DO NOT QUOTE OR CITE
-------
75
Control
Single (450 kg N/ha) Fertilization
APHJS JUt.ZJ' OCT2B FEBI
1987
APH4 MAV2
1SBS
Multiple (37.5 x 3 kg K/ha) FertHtzation
APR28 JUUS7 OCTSB FS1
AW * MAY 2
1 9BB
Figure 10-7. Soil solution nitrate concentrations in untreated (top), single (450 kg N
ha"1, center), and multiple (37.5 kg urea-N ha"1, 3 tunes, bottom)
applications of urea-N. (After Tschaplinski et al., in press.)
August 1991
10-52 DRAFT-DO NOT QUOTE OR CITE
-------
p
August 1991
10-53
DRAFT-DO NOT QUOTE OR CITE
-------
1 soil, and the total ecosystem increase with the rate of fertilization and show no sign of
2 levelling off even at rates of fertilizer N input of up to 1,500 kg/ha (Figures 10-8 to 10-10).
to
JC
c
Ci
*•»
c
o
a>
CC
Ecosystem N Retention vs Fertilizer N Input
2000
1000-
y « 58.682 + 0.51747x R*2 = 0.592
s . •
I " " —— —'
1000 2000
Fertilizer Input (kg/ha)
Figure 10-8. Ecosystem recovery of fertilizer N as a function of fertilizer N input.
1
2
3
4
Table 10-13 gives a summary of N budgets from the nutrient cycling literature and from
the recently completed Integrated Forest Study (IPS; Johnson and Lindberg, in press). In this
summary, atmospheric inputs are compared with outputs via soil solution or streamwater
(primarily as NO3") and vegetation increment, or the N necessary to build perennial tissues in
August 1991
10-54 DRAFT-DO NOT QUOTE OR CITE
-------
Tree N Retention vs Fertilizer Input
a
"S
tc
u
£
y = - 0.44527 +• 0.26974X R*2 = 0.605
1000
Fertilizer Input (kg/ha)
2000
Figure 10-9. Tree recovery of fertilizer N as a function of fertilizer N input.
1 biomass (bole, branches). It should be noted that the studies prior to IPS measured N
2 deposition principally by bulk precipitation, which substantially underestimates N deposition
3 in many polluted sites (e.g., Lindberg et al., 1986). Most of the IPS data include estimates
4 of both wet and dry deposition, and therefore N deposition values reported there are often
5 much greater than those that would have been reported using bulk collectors. For that
6 reason, the IPS data is shown separately from previous data in Figures 10-11 to 10-13. It
7 should also be noted that vegetation N uptake values in each of these systems are much
August 1991
10-55 DRAFT-DO NOT QUOTE OR CITE
-------
Soil N Retention vs Fertilizer Input
800 T
600-
m
O9
C
Cl
400-
o
to
200-
y = 66.634 + 0.23665X R*2 = 0.241
1000
. 2000
Fertilizer Input (kg/ha)
Figure 10-10. Soil recovery of fertilizer N as a function of fertilizer N input.
1 higher than vegetation increment, since uptake includes N taken up and returned annually via
2 litterfall and foliar leaching. Vegetation increment was chosen for this analysis because it
3 represents the net N demand of growing vegetation which must be satisfied from sources
4 external to the N cycle (atmospheric deposition or soil "mining").
5 The data in Table 10-13 and Figures 10-11 to 10-13 reveal some interesting contrasts
6 between ecosystem retention of fertilizer vs atmospherically-deposited N. First, total
7 ecosystem retention of atmospherically-deposited N ranges from over 99% to -266%, with no
August 1991 10-56 DRAFT-DO NOT QUOTE OR CITE
-------
&Q
t/5
«w
)
60
j?
°°
00 £
-
:a « 8 B
-
1
a,
1
S S S
I
August 1991
IQ-57 DRAFT-DO NOT QUOTE OR CITE
-------
>.
"5
en
e
"c
tr
Ecosystem N Retention vs Atmospheric N Input
30
20-
10-
0-
-10-
-20-
-30'
*•
•O
20
—1—
40
—I—
SO
80
Atmospheric N Input (kg/ha/yr)
Figure 10-11. Ecosystem N retention as a function of atmospheric N input.
1
2
3
4
5
6
7
apparent relationship to atmospheric input (Figure 10-10). Secondly, vegetation N increment
accounts for nearly all ecosystem N retention in most (19 of 24) cases, and calculated soil N
retention is low and frequently negative (14 of 23 cases) (Table 10-13; Figures 10-12 and
10-13). There is no relationship between atmospheric N deposition and either tree increment
or calculated soil retention (Figures 10-12 and 10-13).
The pattern of calculated soil N vs deposition in Figure 10-13 suggests that heterotrophs
are very poor competitors for N, even at very low N input levels. Indeed, it appears as if the
August 1991 10-58 DRAFT-DO NOT QUOTE OR CITE
-------
Tree N Increment vs Atmospheric N Input
80
BO-
C8
c
ffl
E
o
k.
o
c
40-
20
* BULK
O IPS
£2-
10
—1—
20
—T—
30
40
SO
60
Atmospheric Input (kg/ha/yr)
Figure 10-12. Tree N increment as a function of atmospheric N input.
••=„".' - .- ! - - '
1 soil is being "mined" for the N necessary to supply vegetation increment systems with very
2 low atmospheric N inputs. This is readily apparent when N output is plotted as a function of
3 input minus vegetation increment (Figure 10-14). Input minus increment can be thought of
4 as N that is available for either (1) soil heterotroph uptake or (2) nitrate leaching. , A negative
5 value for input-increment implies that either the soil is being "mined" for N to supply tree
6 needs or that there is an unmeasured N input contributing to tree N needs. In either case the
7 data suggest that, contrary to views expressed in the literature (see review above), trees are,
8 in the end, more effective competitors for N than soil heterotrophs. Similarly, the nearly 1:1
August 1991
10-59 DRAFT-DO NOT QUOTE OR CITE
-------
(3
s
c
o
<*•*
DC
Z
O
03
T3
ffl
£
o
§
Calculated Soil N Retention vs N input
20
-20-
-40-
-60
O
6
.Q-
•O
10
20
30
T~
40
—i—
50
60
N Deposition (kg/ha)
Figure 10-13. Calculated soil N retention (Input-increment-leaching) as a function of
atmospheric N input.
1
2
3
4
5
6
7
relationship between N output and input-increment after the latter exceeds 0 (r2 = 0.84)
indicates that N deposited in excess of vegetation needs is not taken up by heterotrophs but
rather is subject to nitrification and nitrate leaching, perhaps because heterotrophs in these
systems are limited by organic substrates or other nutrients.
There are several possible explanations for the rather striking differences in soil
N retention and loss patterns between fertilizer and nutrient cycling/air pollution studies.
Firstly, heterotrophic demand for N in fertilized sites is likely to be greater than in sites
August 1991 10.60 DRAFT-DO NOT QUOTE OR CITE
-------
Nitrogen Leaching vs Inputs-Vegetation Increment
-60
Input Minus Vegetation Increment (kg/ha/yr)
Figure 10-14. N leaching as a function of atmospheric N input minus tree N increment.
Points above the 1:1 line imply net soil loss, and points below the line
imply net soil retention.
1 subjected to chronically elevated atmospheric N inputs. Fertilizer N is typically applied to
2 N-deficient ecosystems where N demand by soil heterotrophs is likely to be high, whereas
3 heterotrophic demand for N may have been substantially satisfied in sites with chronically
4 high atmospheric N inputs. Heterotrophic activity in fertilized sites is also likely to be
5 stimulated by mobilization of soil organic C which typically occurs after fertilization
6 (especially with urea; Ogner, 1972; Foster et.al., 1985a). Secondly, as noted above, the
7 slow, steady inputs of N via air pollution, like slow, steady inputs of fertilizer N probably
August 1991
10-61
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
, 5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
favor nitrification. Thirdly, non-biological retention of N is likely to be greater with
fertilization than atmospheric deposition. Ammonium (NH^."1") and ammonia (NH3) fixation
in 2:1 clays is likely to be substantially increased under conditions of high concentrations of
one or both following fertilization. It has also been shown that NH3 can react chemically
with soil organic matter to form very stable, non-labile compounds (Foster et al., 1985b).
Conditions following urea fertilization are especially conducive to these reactions in that pH
is increased and NH3 concentrations are high. These conditions would not normally occur in
sites subject to chronically high atmospheric N inputs.
10.3.4 Effects of Pollutant N Inputs on Soils
10.3.4.1 Soil Biota
The most obvious and immediate effects of pollutant N inputs on soils are those on the
microbial community. An increased activity of heterotrophs and nitrifiers associated with
increased decomposition and nitrification rates would seem a likely result of increased
N inputs. Studies of microbial responses to N fertilization have produced mixed results,
however. Kelly and Henderson (1978) found increased bacterial activity but reduced
invertebrate populations one year after fairly high levels of urea fertilization (550 and
1,100 kg N/ha). This change was important because invertebrates play a major role in the
initial breakdown of Utter. However, the authors found little effect of fertilization on the
decomposition of white oak leaf litter. Kowalenko et al. (1978) found that fertilization with
ammonium nitrate and potassium chloride caused a reduction in soil microbial activity (as
measured by carbon dioxide evolution) for at least three years. This may have been due to
toxic or shock effects due to very large increases in both nitrogen and other ions over a very
short time. Weetman and Hill (1973) reviewed the effects of fertilization on soil flora and
fauna and concluded that fertilization had a lasting, stimulating effect despite short-term toxic
effects of fertilizer components (especially ammonium). Again, we must consider the effects
of single, large inputs of N typical of fertilization studies as opposed to the slow, steady
inputs of N at lower concentration typical of pollutant inputs. Aside from the limited
information on effects on nitrifiers, virtually nothing is known as to as to the effects of slow,
steady inputs of N on soil microbial communities.
August 1991 '
10-62 DRAFT-DO NOT QTJOTE OR CITE
-------
1 10.3.4.2 Soil Chemistry
2 The foremost concern about long-term, capacity-controlled effects of excessive
r> _|_
3 N deposition and NO3" leaching is soil acidification and the mobilization of Al into soil
4 solution and surface waters. As a prelude to assessing the effects of excessive N deposition
5 on soil acidification and A13+ mobilization, a brief review of the components of soil acidity
6 and cation exchange processes is presented. , ;
7 Soil acidity can be measured in a number of ways, but for the purposes of this
8 discussion, we will refer to base saturation as the primary measure or indicator of soil
9 acidity. Base saturation refers to the degree to which soil cation exchange sites, negatively-
10 charged sites to which positively-charged cations are adsorbed, are occupied with base cations
11 (Ca2+, Mg2+, and K+) as opposed to A13+ and H+. Base saturation is a measure of soil
12 acidification, with lower values being more acid. Figure 10-15 shows a soil with 50% base
13 saturation on the left, and a soil with 10% base saturation on the right.
14 Ulrich (1983) describes the various buffering ranges soils go through as they acidify:
15 first is the base cation buffering range, where incoming acid and base cations are exchanged
16 primarily for base cations with very little H+ and A13+ increase (Figure 10-15, left). As
17 soils acidify, exchangeable, base cations are replaced by exchangeable A13 + and H+, and
18 soils are said to be in the aluminum buffering range (Figure 10-15, right). Incoming cations
19 (acid and base) are exchanged primarily for H+ and A13+ in soils that are in the aluminum
20 buffering range (Figure 10-15, right). . . . •
21 With the use of a simulation model, Reuss (1983) showed that the transition from the
22 base cation to the aluminum buffering range is very abrupt. His results showed that soil
23 acidification has little effect upon the concentration of A13+ in soil solution over a large
24 range of base saturation values above 20%. However, he noted that fairly minor changes in
25 base saturation within the 10 to 20% range can cause quite large increases in soil solution
26 A13+ concentration. This implies that soils with base saturations of 10-20% are extremely
27 sensitive to change (although this does not necessarily imply that vegetation will respond to
28 soil change). A series of simple laboratory column studies could tell us much about how far
29 some of our forest soils are from the aluminum buffering range and how much additional acid
30 input might be required to put them into this range. .
August 1991
10-63
DRAFT-DO NOT QUOTE OR CITE
-------
Input
Input
Mineral
Weathering
Mineral
Weathering
Leachfng
10%
Base
Sat.
Leaching
Figure 10-15. Schematic diagram of cation exchange for base cations, A13+ and H+ in
circumneutral (50% base saturation, left) and acid (10% base saturation,
right) soils.
1
2
3
4
5
6
7
8
9
1
2
3
4
Once soils are in the aluminum buffering range, the rate of base cation leaching will
obviously decrease because A13+ is now a dominant cation in soil solutions. In a soil free of
vegetation, continued inputs from the atmospheric deposition (which contains base cations as
well as H+) will eventually acidify the soil to the point where base cation outputs equal base
cation inputs. With forest or other vegetation growing on the soil, however, continued base
cation uptake could reduce the base saturation of the soil to the point where export of base
cations is less than input by deposition (Figure 10-15, right). Thus, vegetation uptake can,
by depleting soil exchangeable base cations, cause the soil to begin accumulating base cations
even when the soil is subject to high leaching rates. Of course, this accumulation of base
cations is accompanied by substantially increased leaching of A13+, and the potentially
detrimental effects of the latter must be considered.
The same cation exchange principles that will eventually cause a soil to begin
accumulating incoming base cations when soils acidify into the aluminum buffering range can
August 1991 10_64 DRAFT-DO NOT QUOTE OR CITE
-------
1 also cause an ecosystem to begin accumulating an individual cation (Ca2+, Mg2+, or K+) if
2 tree uptake depletes soils of an individual cation (Johnson and Todd, 1987). In this case, the
3 conservation of the individual cation in question need not be accompanied by significant
4 overall soil acidification and increased leaching of A13+; leaching of the other base cations
5 may be increased instead. We noted such a situation with respect to Ca2+ in an oak-hickory
6 forest on Walker Branch watershed, Tennessee (Johnson et al., 1985). In this ecosystem,
7 tree Ca2+ is very high, soils are very low in exchangeable Ca2+, and consequently Ca2+
8 leaching is low. Thus, the ecosystem shows a net Ga2+ gain from atmospheric inputs
9 (accompanied by net losses of Mg2+, K+, and Na+).
10 The greatest uncertainty in assessing and projecting rates of exchangeable base cation
11 depletion and/or soil acidification is the estimation of primary mineral weathering rates. The
12 weathering of primary soil minerals (e.g., hornblende, feldspar, plagioclase) represents an
13 input to the exchangeable base cation pool (Figure 10-16). Calculations of the potential rate
14 of soil change from exchangeable pools and input-output budgets (e.g., Tomlinson, 1983)
15 represent the worst-case scenario; -that is, they assume that, weathering is zero. A high rate of
16 soil leaching offset by a high rate of weathering results in a high rate of turnover but not a
17 net depletion of exchangeable cations.
18 Equations and simple models of soil weathering are available for primary to secondary
19 mineral transformations (e.g., Lindsay, 1979). However, these equations are of little value
20 for soils with sizeable nonexchangeable base cation reserves contained in ill-defined minerals
21 (such as amorphous Fe and Al oxides; Johnson et al., 1985). A further complication arises
22 when mineral weathering is enhanced by organic acids formed in forest litter or exuded by
23 tree roots (Boyle and Voigt, 1973). Thus, at present, there are only empirical approaches to
24 assessing weathering such as mass balance calculations. One mass balance approach involves
25 measuring fluxes and changes in exchangeable cation pools over time and calculating
26 weathering by difference (Matzner, 1983). A simpler mass balance approach is to estimate
27 the total weathering' loss from a soil by the difference in soil element content at present and
28 that of an equivalent amount of primary minerals (i.e., element content at the time the soil
29 began to form) and divide by the amount of time the soil has been exposed to weathering
30 (e.g., since the last glaciation) (Mazzarino et al., 1983). The latter gives an average
31 weathering rate over geologic time, but it does not represent current weathering rates in the
August 1991
10-65
DRAFT-DO NOT QUOTE OR CITE
-------
Low fnput
High Input
Leaching
Mineral
Weathering
Leaching
10%
Base
Sat.
Figure 10-16. Schematic diagram of cation exchange for base cations, A13+ and H+ in
acid soils with low (right) and high (left) atmospheric deposition rates.
1
2
3
4
5
6
7
8
9
10
11
12
13
soil. The former method gives a better estimate of current weathering rates in the soil, but it
is subject to large uncertainties due to errors in each of the estimates used to calculate it.
Nonetheless, the plot-scale mass balance method, while imprecise, seems the best for
obtaining realistic estimates of current soil weathering rates, especially in systems where
leaching has been increased by artificial acid irrigation (e.g., Stuanes, 1980).
Because forest soils acidify naturally, it must be true that weathering rates do not keep
pace with base cation denudation rates, even under pristine conditions. The relative
contribution of acid deposition to the rate of acidification can be assessed by measuring
element fluxes (e.g., Ulrich, 1980; Matzner, 1983; Johnson et al., 1985), and the actual
magnitude of the acidification rate (which equals base cation export minus weathering input)
can be estimated by measuring changes in exchangeable base cations and acidity through time
(taking into account seasonal variations in surface soils; see Haines and Cleveland, 1981).
The effects of excess N and S deposition on the rate of soil acidification cannot be evaluated
August 1991
10-66
DRAFT-DO NOT QUOTE OR CITE
-------
1 by simply measuring changes in soils through time, however, because the natural rate of soil
2 acidification (via natural leaching and vegetation uptake) cannot be accounted for by simply
3 measuring changes in soils. If soils do not change during the measurement period, it can be
4 stated that neither acid deposition nor natural processes have caused soil acidification.
5 However, if soils have acidified, measurements of fluxes are necessary to determine the
6 extent to which acid deposition has contributed to the observed rate of ^acidification.
7 There are very few proven, documented cases in which excessive atmospheric
8 N deposition has caused soil acidification, but there is no doubt that the" potential exists,
9 given high enough inputs for a sufficiently long time. Van Breemen et al. (1982, 1987)
10 report high acidification pressure on forests of The Netherlands subject to very high inputs of
11 N from nearby agricultural activities (often considerably in excess of 50 kg 1ST' ha"Lyr" ;
12 van Breemen et al., 1982, 1987; Nilsson and Grennfelt, 1988). The hydrogen ion budgets
13 for these sites indicate the clear possibility (if not probability) that soils have been acidified,
14 but actual changes in soil acidity over time have not been measured.
15 Soil acidification is usually thought of as an undesirable effect, but in some cases, the
16 benefits of alleviating N deficiency may outweigh the detriments of soil acidification. For
17 instance, Van Miegroet and Cole (1984) found that excessive N2 fixation by red alder (Alms
18 rubrd) caused large increases in NO3" leaching and a significant amount of soil acidification
19 relative to adjacent Douglas-fir (Psmdotusga menziesii) stands, yet Douglas-fir growth is
20 invariably superior on sites formerly occupied by red alder due to the differences in N .status
21 (Tarrant and Miller, 1963; Binkley, 1983; Van Miegroet et al., 1991).
22 .-,-.. . . , . —
23 10.3.5 Effects on Natural Waters
24 A major recent concern over the effects of soil acidification due to atmospheric
25 deposition of both N and S is the mobilization of A13+, which can be toxic to some
26 terrestrial vegetation and might be carried to surface waters where it is toxic to fish. As in
27 the case of soil acidification, a brief review of processes leading to soil solution and surface
28 water acidification will be presented.as a prelude to discussions as to the effects of
29 atmospheric N deposition on these processes. .•
30 Increased concentrations of NO3" or any other mineral acid anion (e.g., SO4 ", or Cl")
31 in soil solution lead to increases in the concentrations of all cations in order to maintain
-n< August 1991
10-67
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
charge balance. Figure 10-16 shows the effects of low (left) and high (right) inputs of
cations (which are also accompanied by low and high inputs of anions, respectively) to'the
fictitious soil with 10% base saturation shown on the right of Figure 10-15. As can readily
be seen, the concentrations of H+ and A13+ in soil solution are determined not only by base
saturation, but also by total cation (and anion) input rates. Extremely acid soils are a
necessary but not sufficient condition for the mobilization of A13+; elevated inputs of cations
and anions, whether by atmospheric deposition, fertilization, or natural processes must also
occur.
The composition of the cations in a solution in equilibrium with soil can be described
fairly accurately by well-known selectivity equations developed more than 50 years ago
(Reuss, 1983). In essence, these equations predict that the concentration of a given cation in
soil solution is governed by the proportion of this cation on the soil exchange complex and
the total ionic concentration in soil solution.
Reuss (1983) points out one very interesting aspect of these equations with respect to
the question of A13+ mobilization: as total ionic concentration increases, the concentration of
A13+ increases to the 3/2 power of the increase in the concentrations of ratio Ca2+ and
Mg2+ and to the third power of K+, Na+, and H+. In other words, as total cation and
anion concentrations increase, individual cation concentrations increase as follows:
A13+ > Ca2+, Mg2+ > K+, Na+, H+. Thus, soil solution A13+ concentrations increase
not only as the soil acidifies (i.e., as the proportion of A13+ on the exchange complex
increases) but also as the total ionic concentration of soil solution increases. (These equations
also imply that K+, Na+, and H+ will be the least affected by increased NO3" leaching.)
There are several studies in which A13+ concentrations in both soil solution and
streamwaters have been shown to be positively correlated with NO3~ concentrations. The
NO3" - A13+ pulses in soil solution have implications for forest nutrition and are invoked in
some hypotheses of forest decline discussed in the next section. Researchers on aquatic
effects of acid deposition have long noted springtime pulses of NO3", A13 + , and H+ in acid-
affected surface waters of the Northeastern U.S. (Galloway et al., 1980; Driscoll et al.,
1989). In less acid systems, NO3" pulses may be associated with base cations rather than
A13+, and H+: Foster et al. (1989) note pulses of NO3" and base cations in soil solutions
and streams at the Turkey Lakes site in Ontario. Driscoll et al. (1989) reviewed the North
August 1991
10-68
DRAFT-DO NOT QUOTE OR CITE
-------
1 American data relevant to the role of nitrogen in the acidification of surface waters and
2 explored relationships between atmospheric N deposition, soil C/N ratio and streamwater
3 nitrate concentrations. They found no consistent relationships between these factors, and
4 suggested that vegetation uptake, as hypothesizedjby Vitousek and Reiners (1975) may be one
5 of the most important factors in determining streamwater nitrate concentrations.
6
7 10.3.6 Effects of Pollutant N Deposition upon Vegetation Nutrient Status
8 Because N is the most commonly limiting nutrient for growth in forest ecosystems in
9 North America (Cole and Rapp, 1981), deposition of N in any biologically available form to
10 most forest ecosystems is likely to produce increased vegetation growth to some extent. The
11 degree of response will depend upon the amount of N deposited, the N demand for
12 vegetation, and the competition from soil heterotrophic organisms for this N, as described
13 above. In addition to changes in growth, increased N deposition can cause significant
14 changes in tree physiological function, susceptibility to insect and disease attack, and even
15 plant community structure. In this section, we will briefly review plant physiological
16 responses related to increased N nutrition (see Section 11.4 for more in depth coverage), and
17 give a more in-depth review of soil-mediated effects of N deposition on vegetation and an
18 update on plant community/successional changes that seem to be occurring in high-deposition
\ f
19 areas of Europe. v
20 , !
21 10.3.6.1 Physiological Effects of Excess N Inputs |
22 Nitrogen addition can have several impacts upon trees in addition to improvement of
23 growth, including susceptibility to other pollutants. Nitrogen fertilization has been noted to
24 increase the resistance of eastern white pine (Pinus strobus) to SO2 injury (Cotrufo and
25 Berry, 1970). Nitrogen fertilization usually depressed mycorrhizal development (Weetman
26 and Hill, 1973; Menge et al., 1977). Because the mycorrhizal association is thought to be an
27 adaption to nutrient deficient conditions, suppression of mycorrhizae by N inputs might be
28 expected.
29 Several hypotheses posed to explain current forest declines in eastern North America
30 invoke the effects of excess N deposition upon physiological processes. These physiological
31 responses generally invoke altered carbohydrate allocation causing increased sensitivity to
August 1991
10-69
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
drought, frost, or insect attack. Friedland, et al. (1984) posed the hypothesis that excessive
N deposition induced growth later into autumn which caused susceptibility to frost in red
spruce in the Northeastern United States, Evans (1986) followed up on this, observing that
winter injury apparently occurred to first-year twigs and adding the alternative hypothesis that
excessive N deposition could have caused reduced bark formation as well as or instead of late
growth into the autumn in first-year twigs. Waring (1987) poses an hypothesis in which
boreal coniferous species are unable, to store nitrate taken up from soil solutions, necessitating
the formation of amino acids in green leaves, causing reduced allocation of carbohydrate to
roots and increased susceptibility to drought and pathogens.
More recent studies on red spruce response to nitrogen lend no support to the various
hypotheses for N-induced physiological damage and decline described above. Sheppard et al.
(1989) found the evidence for pollutant-induced susceptibility to freezing injury in red spruce
to be weak, based upon laboratory studies with detached shoots. DeHayes et al. (1989)
found that treatment of red spruce seedlings with ammonium nitrate increased rather than
decreased cold tolerance. Thus, the hypothesis that nitrogen causes direct damage to red
spruce is not supported by laboratory studies. Climate is thought to play a major role in the
severe red spruce decline in the northeastern U.S., perhaps with some additional exacerbation
due to the direct effects of acid mist on foliage (Lucier et al., 1990). There is some evidence
to suggest that indirect effects of nitrogen saturation, namely nitrate and aluminum leaching,
may be contributing factors to red spruce decline in the southern Appalachians, and this
literature is reviewed below.
10.3.6.2 Soil-Mediated Effects on Vegetation
N inputs in excess of tree and heterotrophic N demand may cause immobilization of
some nutrients (especially P and S) and losses of other cation nutrients due to increased
nitrate leaching, as discussed above. In some cases, the benefits of enhanced N status will
greatiy outweigh the detrimental effects of decreased availability of other nutrients. For
instance, the benefits of N fixation during a red alder (Alnus rubra) stage to subsequent
Douglas-fir (Pseudotsuga menziesii) forests in the Pacific Northwest are well-documented
despite the fact that excessive N fixation during the red alder stage causes considerable
P immobilization and soil acidification (Van Miegroet and Cole, 1984). In other cases,
August 1991
10-70
DRAFT-DO NOT QUOTE OR CITE
-------
1 effects of excessive N deposition may be clearly deleterious to plant nutrition. Boxmann
2 et al. (1988) report that excessive NH4+ deposition to soils in which nitrification is inhibited
3 causes serious nutritional imbalances and even toxic effects to some forests in The
4 Netherlands. According to Boxmann et al. physiological mechanisms for these effects might
5 include inhibition of photosynthetic phosphorylation, starch synthesis, protein synthesis
6 (causing a buildup of amino acids), chlorophyll synthesis, and saturation of membrane lipids
7 with NH4+, reducing their ion selectivity and making them more permeable. Increased
7(2 J_
8 membrane permeability may allow potentially toxic ions to be taken up (e.g., AF"1") and
9 allow nutrient ions to be released (e.g., Ca2+). Deleterious effects of excess N deposition
10 can occur via soil interactions as described above, or via aboveground processes. For
11 instance, Roelofs et al. (1987) report that K and Mg deficiencies in declining Dutch forests
12 are caused by excessive foliar leaching due to high inputs of NH4+.
13 Ulrich (1983) hypothesized that these nitrate-induced Al3"1" pulses during warm dry
14 years caused root damage and were a major contributor to forest decline observed in
15 Germany during the mid 1980s. This hypothesis is disputed by other German forest scientists
16 " who point out that forest decline occurred on base-rich as well as base-poor soils (the base
17 -rich soils not being subject to A13+ pulses) (e.g., Rehfuess, 1987). Mulder et al. (1987)
18 document NO3" - A13 + pulses in soil solutions from forest sites in the Netherlands.
19 Aluminum toxicity is one of several nitrogen-related hypotheses posed to explain forest
20 decline in that country. (Other hypotheses are discussed in the following section.) Johnson
21 et al. (in press) found pulses of NO3" and total Al in soil solutions during late autumn from
22 red spruce forests in the Great Smoky Mountains of North Carolina. The pulses were
23 attributed to a combination of high rates of nitrogen mineralization and low uptake in these
24 over mature forests. The soils at these sites were very rich in N, (up to 10,000 kg N/ha) arid
25 atmospheric N deposition was also quite high (26 kg N ha'1 yr"1), both of which contribute to
26 the high rates of NO3~ leaching at these sites. The peak total Al concentrations (70 ^cM/L)
27 associated with these NO3~ pulses were below the threshold for monomeric Al where visible
28 injury to red spruce seedlings occurs in laboratory studies (200 jwM/L; Joslin and Wolfe,
29 1988), and there was no visible evidence of red spruce decline at these sites. However, the
30 possibility of Al inhibition of Ca and Mg uptake cannot be excluded; spot checks revealed
31 that 80-90% of total Al in these soil solutions was in monomeric form, and inhibition of Ca
August 1991
10-71
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
and Mg uptake at monomeric Al levels of well below 200 /*M/L has been noted (Thornton
et al., 1987). In this vein, it is noteworthy that Bondietti et al. (1989) found an inverse
correlation between Al and Ca concentrations in tree rings of red spruce in the southern
Appalachians.
Shortle and Smith (1988) present an hypothesis for the decline of red spruce in which
Al inhibits Ca uptake, Ca deficiency reduces cambial growth (since the demand for Ca per
unit of cambium surface is constant), reduced cambial growth causes a reduction in
functioning sapwood, and reduced sapwood causes a reduction in leaf area. However, A.H.
Johnson (1983) finds no support for the Al hypothesis in the seriously declining forests of
Camel's Hump, Vermont. He found that while the degree of decline increases with
elevation, both Al concentration and Al:Ca ratios in fine roots decrease with elevation while
the degree of dieback and decline increases with elevation. He further points out that high
elevation soils where much of the decline occurs are histosols (organic soils) where Al
toxicity is unlikely due to the mitigating effects of organics on soil solution Al activity,
Thus, the situation with respect to the Al hypothesis and red spruce decline remains
very unclear. There is little support for the Al hypothesis in the northeast, where decline is
very severe. Cook and Johnson (1989) conclude from extensive tree ring and climatic
analyses that red spruce has been out of equilibrium with its climate for the last 150 years,
making it susceptible to damage from a variety of causes, both naturally- and
anthropogenically-induced. Given the soil solution Al levels found in southern Appalachian
red spruce forests, the possibility of some Al effect cannot be excluded, yet decline in this
region is much more subtle (being evidenced primarily by somewhat controversial tree ring
analyses) and no unexpected levels of mortality have yet occurred.
10.3.6.3 Ecosystem-Level Responses to N Deposition
Growth responses to increased N inputs may not always be regarded as desirable,
especially if they result in changes in species composition. For instance, improved growth
and vitality due to increased N deposition may not be deemed desirable in wilderness areas.
Different genera and species respond differentially to increased N availability; for instance,
deciduous species (aniosperms) generally have a greater demand for N per unit biomass
produced than do coniferous species (gymnosperms) (Cole and Rapp, 1981). Tilman (1987)
August 1991
10-72
DRAFT-DO NOT QUOTE OR CITE
-------
1 found marked changes in species composition as a result of experimental N additions to
2 abandoned old fields in Minnesota. Thus, there is a real possibility for changes in ecosystem
3 composition with increased N loading. Changes from heathland to grassland in Holland have
4 been attributed to high rates of N deposition (Roelofs et al., 1987). Ellenberg (1987) points
5 to further species changes in Central European ecosystems as a likely consequence of elevated
6 N. He states that "more than 50% of the plant species in Central Europe can only compete
7 on stands that are deficient in nitrogen supply".
8 There may be significant ecosystem-level effects of N via host-pathogen interactions.
9 Increased N inputs can affect tree resistance to insect and disease either positively or
10 negatively. Nitrogenous fertilizers are known to reduce the production of phenols in plant
11 tissues, thereby reducing resistance to infection by pathogenic fungi (Shigo, 1973). Hollis
12 et al. (1975) noted that additions of P and N to sites deficient in these elements increased the
13 incidence of fusiform rust in slash pine. On the other hand, increased N input will increase
14 resistance to bark beetle and other insect attacks if it improves tree nutrient status (Weetman
15 and Hill, 1973). In addition to changes in tree physiology, increased N inputs produces
16 changes in stand structure which produce changes in understory composition and microclimate
17 that could either increase or decrease the likelihood of insect or disease attack. Brunsting and
18 Heil (1985), addressing the recent changes from heather (Calluna) to grasses in The
19 Netherlands, note that N fertilization leads to increased growth of grasses only when Calluna
20 stands are opened up by beetle attacks. By increasing the N concentration of heather foliage,
21 high N input stimulates larval growth and increases body weight of beetles.
22 The effects of increased N inputs on host-pathogen interactions remain largely
23 speculative; insufficient research on this subject has been done to make many definitive
24 statements. Nonetheless, these interactions are potentially very important, given the
25 devastation that pathogens can produce, and further attention should be given to the issue of
26 effects of increased N deposition, both positive and negative, on host-pathogen interactions.
27
28 10.3.7 Critical Loads for Atmospheric N Deposition
29 Recently, there have been efforts to set critical loads for N deposition for natural
30 ecosystems (Nilsson and Grennfelt, 1988; Fox et al., 1989; Schulze et al., 1989). In that the
August 1991
10-73
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
values for these critical loads may take on considerable political importance, it is appropriate
to examine the assumptions that have been made in defining them.
The Workshop held at Skokloster, Sweden in March 1988 (Nilsson and Grennfelt,
1988) adopted the following definition for a critical load: "A quantitative estimate of an
exposure to one or more pollutants below which significant harmful effects on specified
sensitive elements of the environment do not occur according to present knowledge". In this
document (Nilsson and Grennfelt, 1988) and the subsequent publication synthesizing much of
it (Shulze et al., 1989), nitrogen critical loads were aimed "to protect soils from long-term
chemical changes with respect to base saturation" (p. 17, Nilsson and Grennfelt, 1988;
p. 451, Schulze et al., 1989). The critical loads for N are estimated from two equations.
The first equation is posed as a one that must be satisfied in order to maintain a constant
exchangeable base cation pool in the soil: ,
BC leaching < BC weathering + BC deposition - BC growth
where BC=base cations. Equation (1) is perhaps best understood by rearranging:
BC leaching + BC growth < BC weathering + BC deposition
(1)
(2)
Equation (2) is simply a statement of mass balance for the soil cation exchange complex
and states that removal rates via leaching (BC leaching) and plant uptake (BC growth) must
be equalled or exceeded by inputs via deposition and weathering (the release of base cations
from unavailable, mineral forms to ionic states available for plant uptake, leaching, or
replenishing cation exchange sites) in order to keep soils from acidifying (keep base
saturation constant). This is followed by another equation describing the roles of NQ3" and
SO42" in causing soil leaching:
Nitrate leaching + Sulfate leaching < BC leaching
(3)
The authors state that equation (3) assumes that all base cation leaching is caused by
nitrate and sulfate, ignoring the potentially substantial cation leaching by naturally-produced
August 1991
10-74
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
carbonic and organic acids (e.g., Johnson et al., 1977). However, the use of the "less than"
(<) as well as the "equal to" (=) sign in Equation (3) does, in fact, allow for leaching by
naturally-produced carbonic and organic acids i Base cation leaching will be less than nitrate
plus sulfate leaching if aluminum and hydrogen ions are present to significant extent in soil
solutions.
Combining (1) and (3), the authors obtain:
Acceptable Nitrate Leaching <, BC weathering + BC deposition
- BC growth - Sulfate leaching
(4)
In obtaining Equation (4), the authors assumed (without stating so) that only the "equal
to" (=) and not the "less than" sign in Equation (3) applied; in short, they assumed that all
base cation leaching was due to nitrate + sulfate leaching, and that no H+ and A13+ leaching
occurred.
To estimate nitrate leaching, the authors use the nitrogen balance equation:
N input < N growth + N immobilization — N mineralization
+ N denitrification — N fixation + N leaching
Again, this equation is best understood by rearranging:
N leaching > (N input + N fixation + N mineralization)
— (N growth + N immobilization + N denitrification)
(5)
(6)
Equation (6) can be thought of as a mass balance equation for the soil inorganic N pool
with the first three terms being inputs to that pool and the second three terms being outputs
from that pool other than leaching. The inputs consist of atmospheric deposition (N input),
fixation (N fixation), and release from soil organic matter during decomposition
(N mineralization). The non-leaching outputs include plant uptake (N growth), heterotrophic
uptake (N immobilization), and denitrification (N denitrification). The remainder must be
leaching (N leaching). It is assumed in their analysis that N denitrification and N fixation are
negligible in forest ecosystems and that N immobilization — N mineralization, which is the
August 1991
10-75
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
net annual N accumulation in the soil, equals only 1-3 kg ISP ha"1 • yr"1. The latter numbers,
are based upon an estimate of the net N accumulation in soils of Sweden since the last
glaciation (obtained by dividing nominal soil N content values by the number of years since
glaciation). Soil N accumulation rates can be much higher: Jenkinson (1970) documents net
annual soil N accumulations of over 50 kg • ha"1 • yr"1 over an 81-year period (from 1883 to
1964) after a former agricultural site (Broadbalk) was allowed to revert to forest at the
Rothamsted Experiment Station in England. This high rate of soil N accumulation was
greater than thought possible from atmospheric deposition alone and may have been in part
due to the action of free-living N-fixers in the soil. Liming may have played some role in
stimulating these high accumulation rates; a nearby site (Geescroft) that had not been limed
showed N accumulations of only 23 kg • ha"1 • yr"1 over the same period (Jenkinson, 1970).
Given these equations and estimates of the various parameters within them, the authors
calculate critical loads for various forest ecosystems. These values range from a low of
3-5 kg N • ha"1 • yr"1 for raised bogs to a high of 5-20 kg N • ha"1 • yr"1 for deciduous
forests. A critical concentration for nitrate in groundwater (10 mg N/L) is then calculated
based upon an assumption of precipitation surplus (precipitation minus evapotranspiration) of
100 to 400 mm yr"1, giving values of 10 to 40 kg N • ha"1 • yr"1.
In contrast to the rather quantitative approach taken at the Skokloster Workshop, a far
more subjective approach is taken in determining critical N loads for wilderness areas in the
U.S. Forest Service-sponsored workshop held at Gary Arboretum, Millbrook, New York in
May, 1988. In this case, rather than attempting to come up with specific critical loads, the
workshop participants were asked to establish "green" and "red" lines, the former being
values below which deleterious effects are very unlikely to occur, and the latter being values
above which deleterious effects will very likely occur. The "Rationale used in selecting
nitrogen values" for terrestrial ecosystem critical loads consists of a brief overview of the
nitrogen cycle and some educated guesswork, in view of the fact that "data on N cycling in
wilderness areas is quite scarce at best, and in many areas completely lacking" (p. 12).
Despite the lack of N cycling data, the authors provide guesses at "green" and "red" line
values for specific wilderness areas ranging from 3-10 kg N kg • ha"1 • yr"1 for "green"
values and 10-15 kg • ha"1 • yr"1 for "red" values. These values quantitatively similar to
August 1991
10-76
DRAFT-DO NOT QUOTE OR CITE
-------
1 those obtained in the Skokloster workshop, and actually show very little spread between
2 "green" and "red" lines.
3 :
4 10.3.8 An Evaluation of Critical Loads Calculations for N Deposition
5 There are a number of points that need to be emphasized before the Skokloster critical
6 load values are used for assessment or policy-making. First, the assumption that soils can
7 accumulate only 1-3 kg N • ha"1 • yr"1 is certainly not valid over the short term in most forest
8 ecosystems, as shown amply by a number of forest fertilization studies described in
9 Section 10.3.3. Having stated that, however, it should also be noted that both heterotroph
10 and ecosystem-level recovery of atmospherically-deposited N seems to be lower than that of
11 fertilizer N, as also noted also in Section 10.3.3. The authors of the critical load document
12 (Nilsson and Grennfelt, 1988) recognize that N retention in the soil can be quite high on a
13 temporary basis, but they assume that only net increment in trees is significant over the
14 longer-term (i.e., harvest rotation lengths of 50-100 years). Nonetheless, even "temporary"
15 retention of atmospherically-deposited N could be significant: if N-deficient systems can
16 retain as much as 600 kg N • ha"1 in the soil by heterotrophs (see Table 10-13), an
17 atmospheric N input of 25 kg '• ha"1 ' yr"1 could be retained for 24 years. Recall that
18 Jenkinson (1970) found an average annual N accumulation of about 25 kg • ha"1 • yr"1 in soils
19 at the Rothamsted Experiment Station in England over an 80-year period (1888-1962). This
20 accumulation, which was calculated by differences in measured soil N content over time, is
21 of special interest in that it actually exceeded estimated atmospheric N deposition over that
22 period. It seems clear that estimates of atmospheric N inputs to these sites are low, due
23 either to underestimates of dry deposition or N-fixation
24 A critical unknown in soil heterotrophic N retention is the change (if any) in the relative
25 competitiveness of trees, heterotrophs, and nitrifiers, as noted earlier. There is some
26 evidence to suggest that nitrifiers become more competitive with slow, steady inputs (Johnson
27 and Todd, 1988). Also, it is clear that tree N from the irrigation and fertilizer experiments
28 noted above (Aronsson and Elowson, 1980; Ingestad, 1980; Landsberg, 1986) can increase
29 substantially with increasing N deposition rate, bringing into question calculations of N
30 sequestering by trees from areas that are not N-saturated.
August 1991
10-77
DRAFT-DO NOT QUOTE OR CITE
-------
1 Also inherent in at least the final calculations is the assumption that no natural leaching
2 processes are currently contributing to soil acidification. That is, all base cation leaching is
3 attributed to sulfate and nitrate. This assumption is clearly false; carbonic and organic acids
4 are present in all soil systems and contribute to leaching and acidifying processes to varying
5 degrees (Johnson et al., 1977; Richter et al., 1983; Uirich, 1980). The net result of this
6 assumption, ironically enough, is to underestimate soil acidification (i.e., the acidification by '
7 carbonic and organic acids do not enter into the calculations) and therefore set critical loads
8 (as defined in these calculations) too low. -
9 The weakest link in this chain of calculations is, as always, BC weathering. While the
10 chemical transformations of many weathering reactions are well-known (Lindsay, 1979), :
11 quantification of weathering rates under field conditions has remained elusive. The
12 weathering numbers used in calculating these critical loads are crude mass balance estimates
13 based amounts of minerals and cation nutrients left in soils 8,000-12,000 years after the last
14 glaciation (when fresh minerals were first exposed). These calculations do not account for
15 changes in weathering rates with time (rates were likely much faster initially with fresh
16 minerals than later during the course of weathering) nor do they account for the possibility of
17 increased weathering rates with increased acidification pressure or with vegetation rooting
18 (e.g., Boyle and Voigt, 1973).
19 The entire critical loads concept which formed the basis of the Skokloster document is
20 based upon preventing soil acidification. Implicit in this goal is the assumption that soils
21 reach and remain in some kind of steady-state, non-acid condition in nature, an assumption
22 that is probably fallacious given the presence of extremely acid soils in pristine, unmanaged
23 forests (e.g., Johnson et al., 1977). Furthermore, it is not at all clear that soil acidification .is
24 always harmful. As shown in the red alder/Douglas fir succession example above, the
25 benefits of N deposition may well outweigh the detriments of soil acidification. It should be
26 kept in mind tihat forests of the northern hemisphere have historically been nitrogen deficient,
27 and that growth increases brought about by fertilization (often at levels far in excess of
28 critical loads) have been regarded as beneficial, at least in commercial forest lands. Value
29 judgements inevitably come into play in setting critical loads for pollutant deposition of
30 nutrients, especially in the case of N.
August 1991
10-78
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
The "green" and "red" lines for N deposition established for wilderness areas in the
Gary Arboretum workshop (Fox et al., 1989) were almost totally subjective guesses and are
therefore open to many criticisms and argument. Given the fact that wilderness areas,
especially those in the western U.S., are very likely nitrogen-limited, even the "green" lines
are not a guarantee of having no effect, as is acknowledged by the authors. They state,
however, that "in our judgement, the Green Line levels are sufficiently low that perceptible
deleterious effects upon plant health, changes in species composition, or degradation of water
quality are unlikely". In view of the very low N deposition rates in some parts of the
western U.S., (1-2 kg • ha"1 • yr"1;,Table 10-13), it seems likely that increases of up to
10 kg • ha"1 • yr"1 will result in some increases in plant growth, plant health, and, quite
possibly, changes in species composition. The judgement that deleterious effects on plant
health and water quality are unlikely to occur at these levels seems to be a reasonable one for
the short term (i.e., until biological N demand is satisfied in these slow-growing ecosystems),
but remain open to serious question over:the long term.
10.3,9 Conclusions
There is little doubt that N deposition has a pronounced effect on many, if not most
terrestrial ecosystems. Because most forest ecosystems in North America are'N deficient,
one of the most noticeable initial changes in response to increased N deposition is likely to be
a, growth increase. Whether this growth increase is deemed desirable or undesirable in a
particular ecosystem is entirely a matter of management objectives (timber production or
species preservation), and, ultimately, value judgements by society.
All current information indicates that such "N-saturated" forests are relatively rare and
limited in extent (e.g., Cole and Rapp, 1981), especially in managed forests. Forest
management practices, especially with respect to harvesting and fire, will have a major effect
upon the degree to which forests become N saturated. The critical load values given in the
Skolster document (Nilsson and Grennfelt, 1988) are unlikely to produce N-saturation in
highly-productive, intensively-managed forests of the timberbelts in the southeastern and
northwestern U.S. that are frequently harvested and/or subjected to control burning. Indeed,
there is considerable concern that intensive management practices in these forests are causing
N depletion (Boyle and Ek, 1972).
August 1991
10-79
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
Given the great variation in both natural forest N uptake rates and management
intensity, it is not reasonable to assign one critical load for all forest ecosystems. Intensively
managed, short-rotation forests might beneficially utilize up to 100 kg N ha"1 yr"1, whereas a
value as low as 10 kg N ha"1 yr"1 may be produce undesired growth increases in very slow-
growing virgin forests in wilderness areas. Given some knowledge about forest uptake rates
and current N status, critical loads might be calculated on a site-by-site basis, but regional
values will invariably prove invalid for many (if not most) forests within the region to which
they are applied. ' !
10.4 TERRESTRIAL ECOSYSTEM EFFECTS-VEGETATION
Subsequent to the dry or wet deposition of N forms from the atmosphere (Section 10.2)
nitrogen containing compounds can impact the terrestrial ecosystem through direct effects on
plant metabolic processes, or indirectly by modifying the nitrogen cycle and associated soil
chemical properties. However, interpretation of the effects of N deposition at the level of the
ecosystem becomes difficult because of the complex interactions which exist between
biological, physicochemical, and climatic factors (U.S. Environmental Protection Agency,
1982).
10.4.1 Direct Effects
Direct effects of reactive nitrogen compounds on terrestrial ecosystems are defined as
those effects that impact individual plants by disturbing "normal" physiological processes.
Because information on the direct effects of NO and NO2 alone and in combination with
other pollutants have been described in detail in Sections 9.3 through 9.6, they will not be
discussed here. :
Very little information is available on the direct effects of nitric acid vapor on
vegetation and essentially no information on its effects on ecosystems. Norby et al. (1989)
reported that nitric acid vapor (0.075 ppmv) induced nitrate reductase activity (NRA) in red
Spruce foliage. Because the induction of NRA is a step in the process leading to the
formation of organic nitrogen compounds (amino acids), the nitrate from nitric acid could
function as an alternative source of nitrogen for plant growth. However, in plants under
August 1991
10-80
DRAFT-DO NOT QUOTE OR CITE
-------
1 stress the reduction of nitrate to amino acids consumes energy needed for alternative
2 metabolic processes; a potentially slight negative impact.
3 The effects of ammonia, a reduced nitrogen gas, have been summarized by Van Der
4 Eerden (1982), however, ammonia concentrations seldom reach phytotoxic levels in the
5 'United States, consequently it will not be extensively discussed here (U,S. Environmental
6 Protection Agency, 1982). In contrast, high ammonia concentrations in Europe have been
7 observed (van Dijk and Roelofs, 1988). Van Der Eerden (1982) summarized available
8 information on the response of crop and tree species to ammonia fumigation and concluded
9 that the following concentrations produced no adverse effects:
10 0.107 ppmv (75/zg m~3), yearly average
11 0.858 ppmv (600/ig m"3), daily average
12 14.3 ppmv (10,000 /ig m"3) hourly average.
13
14 Submicron, ammonium sulfate aerosols have been shown to affect foliage of Phaseolus
15 vulgaris L. (Gmur et al., 1983). At a concentration of 26 mg m"3 (37 ppmv), three weeks of
16 exposures produced leaf chlorosis, necrosis and loss of turgor. Gmur et al. (1983) reported
17 that these foliar symptoms were not correlated with changes in shoot or root dry mass, and,
18 suggested that no relationship to plant growth was expected. However, the 3-week
19 experiment was not long enough for significant changes in dry matter to be observed. The
20 level of ammonia producing'the leaf effects (37 ppmv) exceeds normal ambient levels for the
21 U.S., but it is representative of reported high concentration episodes in Europe (Gmur et al,,
22 1983). Cowling and Lockyer (1981) reported beneficial effects of ammonia on the growth of
23 Lolium perenne L. due to sorption of NH3 .nitrogen through leaves. Van Hove et al. (1989b)
24 studied the effects of 50 and 100 /ig m"3 NH3 on Populus euramericana L. over a 6 to
25 8-week period and found increases in photosynthesis at 100 /ig m"3, but no 'changes in
26 stomatal characteristics up to that level of NH3.
27 . , ......
28 10.4.2 Indirect Effects
29 Indirect effects of dry nitrogen deposition to terrestrial ecosystems result from the
30 addition of mtrogen to ecosystems at a rate above that experienced during normal successional
31 processes leading to ecosystem eutrophication. Positive responses to added nitrogen would be
32 anticipated in many cases because many natural systems are nitrogen limited (Krause, 1988;
August 1991
10-81 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
National Research Council, 1979; see also Sections 10.3 and 10.5). However, if atmospheric
additions of nitrogen exceed the "buffering" capacity of an ecosystem, alterations in soil
chemistry are expected to take place (see Section 10.3). Inputs of nitrogen to natural systems
are hypothesized to alleviate deficiencies and allow increased growth of some plants, but in
doing so may also impact interplant competitive relationships which would result in altered
species composition and diversity 'in sensitive ecosystems (U.S. Environmental Protection
Agency, 1982; Ellenberg, 1987). Schulze (1989) has also proposed that excessive additions
of nitrogen lead to nutrient deficiencies of other,elements,(Ca, Mg). Aber et al. (1989)
outlined a hypothetical progression of the effects of excessive nitrogen additions for northern
forest ecosystems, and concluded that these systems have a limited capacity to accumulate
nitrogen.
In addition to the potential for increasing plant productivity through fertilization, the
deposition of nitrogen from the atmosphere to ecosystems has been hypothesized to disrupt
normal nutrient cycles and physiological processes, resulting in increased susceptibility of
forests to other environmental stresses (Lindberg et al., 1987; Nihlgard, 1985; McLaughlin,
1985; Schulze, 1989). Physiological imbalances resulting from excessive nitrogen additions
are also hypothesized to disrupt the winter hardening process (Nihlg&rd, 1985; Friedland ,
etal., 1984; Waring, 1987), produce nutrient imbalances (Nihlgard, 1985; Waring, 1987;
Schulze, 1989), and altered carbon allocation patterns within plants (Nihlgard, 1985;
McLaughlin, 1985). Altered shoot:root ratios resulting from different patterns of carbon
allocation can lead to increased susceptibility to drought because shoots grow^at the expense
of roots under high nitrogen availability (Freer-Smith, 1988; Norby et al., 1989;
McLaughlin, 1985; Waring, 1987). Changes in carbon:nitrogen ratios of tissues resulting
from an excessive supply of nitrogen can also result in altered host-pathogen, mycorrhizal,
and pest-plant interactions (Grennfelt and Hultberg, 1986; Nihlgard, 1985). In addition to
these indirect soil-mediated effects on individual plants, Ellenberg (1987) has suggested that
current balances of interspecific competition in some sensitive ecosystems might be altered by
additional sources of nitrogen and result in the displacement of existing species by plants that
can utilize the excess nitrogen more efficiently. For example, Roelofs et al. (1987) proposed
that ammonia/ammonium deposition leads to heathland changes via two modes
(1) acidification of the soil and associated loss of cations such as K+, Ca2+, and Mg2+
August 1991
10-82
DRAFT-DO NOT QUOTE OR CITE
-------
1 (2) nitrogen enrichment leading to "abnormal" plant growth rates and altered competitive
2 relationships.
3 Although much has been hypothesized about the impact of excessive inputs of nitrogen
4 into forest ecosystems, direct experimental information to prove or disprove these hypotheses
5 is not widely available. Margolis and Waring (1986) showed that fertilization of Douglas fir
6 with nitrogen could lengthen the growing season to the point where frost damage became a
7 .problem. However, Klein and Perkins (1987) presented other evidence which showed no
8 additional winter injury of high elevation conifer forests when fertilized with 40 kg total
9 nitrogen ha'1 yr"1. Van Dijk et al. (1990) conducted a greenhouse study to determine the
10 impact of ammonium in rainwater on three coniferous trees (Douglas fir, Gorsican pine, and
11 Scots pine) and found no sign of deterioration in seedlings receiving nitrogen at the rate of
12 48 kg ha"1 yr'1. At the very high rates of'application of 480, kg N ha'1 yr"1 increases in
13 shoot/root ratio, and reductions in fine root and mycorrhizal biomass were observed.
14 However, this level of nitrogen addition (i.e., simulated deposition) is approximately one
15 order of magnitude greater than most rates of deposition in North America or Europe. Kenk
16 and Fischer (1988) summarized fertilization experiments on German forests and found little
17 evidence for negative effects, but some indication of increased growth since 1960 that could
18 be the result of atmospheric N deposition was indicated for Norway spruce. Miller and
19 Miller (1988) concluded that fertilizer trials are not be appropriate for extrapolation as
20 indicators of forest response to N deposition (i.e., the timing of applications is typically quite
21 different), but nevertheless they also suggested that results of such trials ought to be
22 reconcilable with the "natural" phenomenon.
23 De Temmerman et al. (1-988) provided data showing increased fungal outbreaks and
24 frost damage on several pines species exposed to very high ammonia deposition rates
25 (> 350 kg ha"1 yr"1). Numbers of species and fruiting bodies 'of fungi have also increased
26 concomitantly with nitrogen deposition in Dutch forests (van Breemen and van Dijk, 1988).
27 An increase in total amino acid concentrations in needles known to take place in response to
28 dry deposition of NOX (Section 10.2), has also been suggested to favor outbreaks of insect
29 pests (Waring and Pitman, 1985; White, 1984). Schulze (1989) presents a clear progression
30 of evidence which indicates that canopy uptake of nitrogen together with root uptake has
31 caused a nitrogen imbalance in Norway spruce leading to forest decline.
August 1991
10-83
DRAFT-DO NOT QUOTE OR CITE
-------
1 Excessive nitrogen inputs to terrestrial ecosystems can cause differential competitive
2 advantage among plants within a heathland (Heil and Bruggink, 1987; Heil et al., 1988).
3 The authors established that the changing nature of unmanaged heathlands in the Netherlands,
4 where Calluna vulagris is being replaced by grass species, is a result of the eutrophic effect
5 of acidic rainfall and large nitrogen inputs arising from intensive farming practices in the
6 region. Both Calluna vulagris and Molinia caerulea are stress tolerant species (Grime, 197
-------
1 in response to added N or (3) added N leads to increased losses of nitrate in streamwater, and
2 concluded that the third was most reasonable (see also Section 10.3). Brown et al. (1988)
3 reported that a recent workshop concluded that N saturation could be best defined as
4 occurring when N outputs from ecosystems exceeded inputs. This conclusion was based on a
5 model of plant/soil N saturation put forth by Agren and Bosatta (1988). Aber et al. (1989)
6 similarly define nitrogen saturation as the availability of ammonium and nitrate in excess of
7 total combined plant and microbial nutritional demands. The concept of N saturation leads to
8 the possibility of defining a critical N load (deposition rate) at which no change or deleterious
9 impacts will occur to an ecosystem (Nilsson, 1986). It is important to recognize that the
10 magnitude of such a "critical load" will be site and species specific being highly dependent on
11 initial soil chemistries and biological growth potentials (i.e., nitrogen demands).
12
13 10.4.3.1 Critical N Loads That Have Been Proposed
14 Skeffmgton and Wilson (1988) summarized and discussed the following possible criteria
15 as potentially useful for defining appropriate critical N loads on ecosystems:
16 • Prevent nitrate levels in drinking or surface waters from rising above
17 standard levels
18 • Ensure proton production less than weathering rate
19 • Maintenance of a fixed ammonia-base cation balance
20 • Maintenance N inputs below N outputs (the N saturation approach)
21 • Minimize accelerations in the rates of ecological succession (vegetation changes
22 due to altered interspecific competition).
23
24 De Vries (1988) has also defined criteria for a combined critical load for nitrogen and
25 sulfur for Dutch forest ecosystems based on the following: N contents of foliage, nitrate
26 concentrations in groundwater, NH4/K ratios, Ca/Al ratios, and Al concentrations in soil
27 solution. Based on these criteria, De Vries concluded that current rates of N and
28 S deposition in the Netherlands exceed acceptable levels.
29 Schulze et al. (1989) have also proposed critical loads for N deposition based on an
30 ecosystem total anion and cation balance. This approach makes the assumption that processes
31 determining ecosystem stability are related to soil acidification and nitrate leaching (see also
32 Section 10.3.6). They concluded that in order to limit the mobilization of aluminum and
33 other heavy metals resulting from acidification and nitrate leaching (a negative result), critical
34 nitrogen deposition rates could not exceed 3-14 kg N ha'1 yr'1 for silicate soils or 3 to 48 kg
August 1991 10-85 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
N ha"1 yr"1 for calcareous based soils. Other critical loads have been proposed at rates of
nitrogen deposition ranging from as little as 1 to levels near 100 Kg N ha"1 yr"1 depending on
the impacts considered acceptable and the criteria used to define the critical load.
Critical loads less than 30 kg ha"1 yr"1 have been proposed based on criteria to minimize
species changes (van Breeman and van Dijk, 1988; Liljelund and Torstensson, 1988). Using
the criteria that ecosystem nitrogen inputs should not exceed outputs, critical loads have been'"'
proposed as low as 1-5 kg N ha"1 yr"1 for poorly productive sites with low productivity or in
the range from 5-30 kg N ha"1 yr"1 for sites having medium quality soils and for common
forested systems (Boxman et al., 1988; Rosen, 1988; Skeffington and Wilson, 1988; World
Health Organization, 1987).
In their summary of a recent conference on critical nitrogen loading, after discussing
various options for setting a critical N load Skeffington and Wilson (1988) concluded that
"we do not understand ecosystems well enough to set a critical load for N deposition in a
completely objective fashion". Brown et al. (1988) further concluded that there was probably
no universal critical load definition that could be applied to all ecosystems, and a combination
of scientific, political, and economic considerations would be required for the application of
the critical load concept.
The following terrestrial ecosystems have been suggested as being at risk'from the
deposition of nitrogen-based compounds:
• heathlands with a high proportion of lichen cover,
• low meadow vegetation types used for extensive grazing and ..
haymaking, and
• coniferous forests, especially those at high altitudes (World Health
Organization, 1987).
These oligotrophic ecosystems are considered at risk from atmospheric nitrogen inputs
because plant species normally restricted by low nutrient concentrations could gain a
competitive advantage, and their growth at the expense of existing species would change the
"normal" species composition and displace some species entirely (Ellenberg, 1987; Waring,
1987). Sensitive natural ecosystems, unlike highly manipulated agricultural systems, may be
prone to damage from exposure to dry deposited nitrogen compounds because processes of
natural selection whereby tolerant individuals survive may not be keeping pace with the
current levels of atmospheric N deposition (World Health Organization, 1987).
August 1991
10-86
DRAFT-DO NOT QUOTE OR CITE
-------
1 10.4.3.2 Current Rates of Total N Deposition
2 Application of the concept of critical N loading has not yet been widely adopted in
3 North America (based on amount of published data), but a comparison of total N deposition
4 data for North America and proposed critical loads just discussed should provide a reasonable
5 comparison of the status of terrestrial systems with respect to changes expected from elevated
6 ..... levels of nitrogen deposition. Tables 10-14, 10-15, and 10-16 summarize information
7 = regarding the total deposition of nitrogen to a variety of ecosystems/forest types in North
8 America. Table 10-14 summarizes information regarding the total deposition of nitrogen to a
9 variety of ecosystems/forest types or regional areas in North America and Europe.
10 Nitrogen deposition can be divided into the four categories depending on its origin:
11 -. cloudwater, precipitation, dry particles, and gaseous forms. Table 10-15 summarizes wet
12 deposited nitrate and ammonium deposition data for various states that were part of the
13 National Acid Deposition Program. Table 10-16 specifically addresses the issue of
14 relationships between ecosystems nitrogen inputs and outputs. Data in these tables indicates
15 that total deposition of nitrogen in North America is typically less than rates found for many
16 areas in Europe. North American sites would appear to have total N deposition rates less
17 than 25 kg N ha"1 yr"1. It is also obvious from these summary tables that much of our
18 information on nitrogen deposition is limited to information on nitrate and ammonium
19 deposition in rainfall. Lindberg et al. (1987) concluded that the lack of data on multiple
20 forms of nitrogen deposition limits our ability to accurately determine current levels of
21 nitrogen loading.
22 Olsen (1989) summarized nitrate and ammonium concentration and wet deposition data
23 for the United States and southern Canada for the period from 1979 through 1986. For
24 1986, the greatest annual rates of ammonium and nitrate deposition were localized in the
25 northeastern United Sates and southern Canada (Olsen, 1989). Peak values were 5 and
26 25 kg ha"1 yr"1 for ammonium and nitrate, respectively. Similar wet deposition data for 1987
27 showed peak deposition rates of 3.5 and-16 kg ha"1 yr"1 for ammonium and nitrate,
28 respectively (National Atmospheric Deposition Program, 1988). Zemba et al. (1988)
29 summarized wet nitrate deposition data from 77 stations located in Eastern North America
30 and found that peak nitrate deposition (>20 kg ha"1 yr"1) occurred between lakes Michigan
31 and Ontario. They also found the temporal pattern of nitrate deposition was quite even
August 1991
10-87
DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-14. MEASUREMENTS OF VARIOUS FORMS OF
ANNUAL N DEPOSITION TO NORTH AMERICAN AND
EUROPEAN ECOSYSTEMS
Forms of N deposition (Kg ha'1)
Site Location/
Vegetation
United States
CA, Chaparral
CA, Sierra Nevada
GA, Loblolly pine
NC, Loblolly pine
NC, Hardwoods
NC, White pine
NC, Red spruce
NH, Deciduous
NH, Deciduous
NY, Red spruce
NY, Mixed deciduous
TN, Mixed deciduous
TN, Oak forest #1
TN, Oak forest #2
TN, Oak forest #1
TN, Oak forest #2
TN, Oak forest
TN, Loblolly pine
WA, Douglas fir
WA, Douglas fir
U.S. Regions
Adirondacks
Midwest
Northeast
Northwest
Southeast
S.E. Appalachians
Canada
Alberta (southern)
British Columbia
Ontario
Ontario (southern)
Wet
Cloud Rain
8.2
._
3.7
8.7
4.8
3.7
8.7 6.2
7
9.3
7.3 6.1
4.8
2.9
3.2
2.9
' - 6.9
6.0
4.5
4.3
2.9
j
6.3
- : 4.2
- 21.7
~ 16.6
- 20.6
- : 4.2
7.3
5.5
3.7
2.3
Dry
Particles
—
—
1.0
2.2
0.5
0.9
3.6
—
—
0.2
0.8
4.1
4.4
4.4
1.3
1.2
1.8
0.6
1.3
—
4.7
2.9
—
—
—
3.1
12.2C
—
—
1.4
Gases
—
—
4.2
4.1
—
2.7
8.6
—
—
2.3
2.5
6.1
4.0
4.0
—
—
3.8
1.4
0.6
—
—
—
—
—
—
—
—
—
—
—
Total
23b
(2)°
9
15
5.3
7
27
(7)
(9)
16
8
13
12
11
8
7
10
9
5
(1)
11
7.r-
22
17
21
7.3
19.5
(5)
(4)
3.7
Reference
Riggan et al. (1985)
Williams and Melack (in press)
Lovett (1991)
Lovett (1991)
Swank and Waide (1988)
Lovett (1991)
Lovett (1991)
Likens et al. (1970)
Likens (1985)
Lovett (1991)
Lovett (1991)
Kelly and Meagher (1986)
Kelly and Meagher (1986)
Kelly and Meagher (1986)
Kelly (1988)
Kelly (1988)
Lindberg et al. (1986)
Lovett (1991)
Lovett (1991)
Henderson and Harris (1975)
Driscoll et al. (1989a)
Driscoll et al. (1989a)
Munger and Eisenreich (1983)
Munger and Eisenreich (1983)
Munger and Eisenreich (1983)
Driscoll et al. (1989a)
Peake and Davidson (1990)
Feller (1987)
Linsey et al. (1987)
Ro et al. (1988)
August 1991
10-88
DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-14 (cont'd). MEASUREMENTS OF VARIOUS FORMS OF
ANNUAL N DEPOSITION TO NORTH AMERICAN AND
EUROPEAN ECOSYSTEMS
Forms of N deposition (Kg ha"1)
Site Location/
Vegetation
Wet
Dry
Cloud Rain Particles Gases Total Reference
Fed. Rep. Germany
Spruce (SE slope)
Spruce (SW slope)
Netherlands
Oak-birch
Deciduous/spruce
Scots pine
Douglas fir
Douglas fir
Norway - ,-
Spruce
United Kingdom
Spruce Forest
Cotton grass moor
16.5
24.3
1.9
0.4
19.3
10.3
8.0
8.0
95.7=
0.7
na
na
0.2
13.5
4.0
16.5 Hantschel et al. (1990)
24.3 Hantschel et al. (1990)
24-56b van Breemen and van Dijk
(1988)
21-42b van Breemen and van Dijk
(1988)
17-64b van Breemen and van Dijk
(1988)
l7-64b van Breemen and van Dijk
(1988)
115 Draaijers et al. (1989)
11.2 Lovett(1991)
3-19b Royal Society (1983)
23.4 Fowler et al. (1989)
12.4 Fowler et al. (1989)
*— Symbolizes data not available or in the case of cloud deposition not present.
"Total nitrogen deposition was based on bulk deposition and throughfall measurements and does include
components of wet and dry deposition.
"Measurements of total deposition data that do not include both a wet and dry estimate probably,underestimate
total nitrogen deposition and are enclosed in parentheses.
dlncludes deposition from gaseous forms.
1 throughout the year (Schwartz, 1989). Wet deposition of ammonium (NH4+) in Europe
2 ranges between 3.5 and 17.3 kg NH4+ ha"1 yr"1 (Buijsman and Erisman, 1987; Heil et al.,
3 1987). Boring et al. (1988) have also published an extensive review of the sources, fates and
4 impacts of nitrogen inputs to terrestrial ecosystems.
5 For an oak-hickory forest in eastern Tennessee, dry deposition made up greater than
6 80% of the total atmospheric deposition of nitrogen ions (Lindberg et al., 1986). Barrie and
August 1991
10-89
DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-15. MEAN ANNUAL WET NITRATE AND AMMONIUM
DEPOSITION TO VARIOUS STATES LOCATED THROUGHOUT
THE UNITED STATES*
Forms of N deposition (Kg ha"1)
Location
Pennsylvania
New York
Ohio
Georgia
Tennessee
Illinois
N. Carolina
Arkansas
Virginia
Florida
Oklahoma
Colorado
Alabama
New Mexico
S. Dakota
Texas
No. of Sites Nitrate
3 10.9
5 9.7
2 7.6
1 6.9
1 6.9
4 6.2
4 6.2
1 5.0
1 5.3
2 4.9
3 4.1
4 4.3
1 3.7
4 3.6
1 2.7
3 3.1
Ammonium Total*
1.3 12.2
1.4 11.1
1.7 9.3
1.1 8.0
0.8 7.7
1.3 7.5
1.1 7.3
1.3 6.3
0.5 5,8
0.6 5.5
1.3 5.4
0.6 4.9
0.6 4.3 ,
0.5 ,4.1
1.3 4.0
0.6 3.7
Reference11
National Atmospheric Deposition Program
(1988)
National Atmospheric Deposition Program
(1988)
National Atmospheric Deposition Program
(1988)
National Atmospheric Deposition Program
(1988)
National Atmospheric Deposition Program
(1988)
National Atmospheric Deposition Program
(1988)
National Atmospheric Deposition Program
(1988)
National Atmospheric Deposition Program
(1988)
National Atmospheric Deposition Program
(1988)
National Atmospheric Deposition Program
(1988)
National Atmospheric Deposition Program
(1988)
Bohm (1991)
National Atmospheric Deposition Program
(1988)
Bohm (1991)
National Atmospheric Deposition Program
(1988)
National Atmospheric Deposition Program
(1988)
August 1991
10-90 DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-15 (cont'd). MEAN ANNUAL WET NITRATE AND AMMONIUM
DEPOSITION TO VARIOUS STATES LOCATED THROUGHOUT
THE UNITED STATES*
Location
California
Washington
Wyoming
Arizona
Utah
Idaho
Oregon
Montana
Arizona
Hawaii
Forms
No. of Sites
5
3
3
1
1
1
4
4
1
1
of N deposition (Kg ha'1)
Nitrate
2.9
2.7
2.5
2.6
2.5
2.3
2.1
1.9
1.0
0.08
Ammonium
0.6
0.3
0.4
0.2
0.3
0.3
0.3
0.4
0.2
0.01
i
Total2
3.5
3.0
2.9
2.8
2.8
2.6
2.4
2.3
1.2
0.1
Reference15
Bohm (1991)
Bohm (1991)
Bohm (1991)
Bohm (1991)
Bohm (1991)
Bohm (1991)
Bohm (1991)
Bohm (1991)
National Atmospheric Deposition Program
(1988)
National Atmospheric Deposition Program
(1988)
The states are presented in order of the greatest annual N deposition.
Total deposition data is for wet deposited forms only and as such represents an underestimate of the total
nitrogen loading received by these geographic areas.
bData from National Atmospheric Deposition Program (1988) are for a single year, and data summarized by
Bohm (1991) are for the period from 1985 through 1988.
1 Sirois (1986) estimated that dry deposition contributed 21 to 30% of total NO3" deposition in
2 eastern Canada. Lovett and Lindberg (1986) also concluded that dry deposition of nitrate is
3 the largest form of inorganic nitrogen deposited to oak-hickory forests in eastern Tennessee.
4 Significant nitrogen inputs from the deposition of nitrogen dioxide have been predicted
5 (Hanson et al., 1989; Hill, 1971; Kelly and Meagher, 1986). Duyzer et al. (1987) has also
6 predicted that dry deposition of ammonia can reach levels as high as 54 kg ha"1 yr"1 in areas
7 of high ambient concentration (0.017 ppmv). Typical values of ammonia deposition in
August 1991
10-91
DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-16.
NITROGEN INPUT/OUTPUT RELATIONSHIPS
FOR SEVERAL ECOSYSTEMS
Site/Vegetation
United States
FL, slash pine
GA, Loblolly pine
ME, spruce
NC, Loblolly pine
NC, Oak-Hickory
NC, Red spruce
NC, White pine
NC, White pine
NH, N. Hardwood
NH, N. Hardwood
NY, Deciduous
NY, Red spruce
OR, Douglas fir
TN, Loblolly pine
TN, Hardwood
TN, Hardwood
TN, Hardwood
TN, Oak forest
TN, Oak forest
TN, Shortleaf/pine
TN, Yellow/poplar
WA, Douglas fir
WA, Douglas fir
WA, Red alder
WA, Silver fir
WI, N.hardwoods
Canada
Ontario (maple)
Fed. Rep. Germany
Norway spruce
Beech
Netherlands
Oak
Oak-Birch
Oak
Mixed deciduous
Inputs
5.9b
9b
7.5b
15b
8.2°
27. lb
8.8C
7.4"
6.5
23.6
8b
15.9b
2.0
8.7b
13. 2b
13.0
8.7
7-8d
11. 5b
8.7
7.7
1.7
4.7b
70b
1.3
5.6
7.8
21.8
21.8
45.
54
56
63
Efflux"
0
0
0
0
3.2
11-20
0.2
0
4.0
17.4
1
3
1.5
0-2
4.4
3.1
1.8
1.25
3.2
1.8
3.5
0.6
0
71
2.7
0.05
18.2
14.9
4.4
22
78 ,
28
68
Reference
Van Miegroet et al. (1991)
Van Miegroet et al. (1991)
Van Miegroet et al. (1991)
Van Miegroet et al. (1991)
Cole and Rapp (1981)
Van Miegroet et al. (1991)
Cole and Rapp (1981)
Van Miegroet et al. (1991)
Bormann et al. (1977)
Likens et al. (1977)
Van Miegroet et al. (1991)
Van Miegroet et al. (1991)
Sollins et al. (1980)
Van Miegroet et al. (1991)
Kelly and Meagher (1986)
Henderson and Harris (1975)
Cole and Rapp (1981)
Kelly (1988)
Kelly and Meagher (1986)
Cole and Rapp (1981)
Cole and Rapp (1981)
Cole and Rapp (1981)
Van Miegroet et al. (1991)
Van Miegroet and Cole (1984)
Turner and Singer (1976)
Pastor and Bockheim (1984)
Foster and Nicolson (1988)
Cole and Rapp (1981)
Cole and Rapp (1981)
van Breemen et al. (1987)
van Breemen et al. (1987)
van Breemen et al. (1987)
van Breemen et al. (1987)
August 1991
10-92
DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-16 (cont'd). NITROGEN INPUT/OUTPUT RELATIONSHIPS
FOR SEVERAL ECOSYSTEMS
Site/Vegetation
Norway
Spruce
Sweden
Coniferous
United Kingdom
Mixed hardwood
U.S.S.R.
Norway spruce
Inputs
11.2b
2.1
5.8
1.1
Efflux"
0
0.6-1
12.6
0.9
Reference
. . , , ...
Van Miegroet et al. (1991)
Rosen (1982)
Cole and Rapp
Cole and Rapp
(1981)
(1981)
"An estimate based on nitrogen losses from the soil profile or from streamflow out of a watershed.
blncludes precipitation, cloud (where appropriate), participate and gaseous forms of nitrogen deposition.
"Includes nitrogen inputs from precipitation and particulate forms of deposition.
dMean of two oak forests in east Tennessee.
1 central Europe and Scandinavia range between 20 and 40 kg ha"1 yr"1 (Grennfelt and
2 Hultberg, 1986).
3 Based on the current rates of N deposition (loading) occuring in North America
4 (Tables 10-14 through 10-16) and the proposed critical N outlined in previous sections
5 (Sections 10.3.6 and 10.4.3.1) one might conclude that current rates of N deposition in North
6 America are sufficient to induce minor changes in some ecosystems (i.e.,' rates of deposition
7 in N. America exceed some of the proposed critical load levels). However, because
8 ecosystems have a variable capacity to buffer changes caused by elevated inputs of nitrogen,
9 it is difficult to make general conclusions about the type and extent of change currently
10 resulting from N deposition in North America. Furthermore, current estimates of total '
11 nitrogen deposition to ecosystems and regions of the United States (Tables 10-14 through
12 10-16) usually do not account for gaseous nitrogen losses from ecosystems (e.g., N2O and
13 NH3), therefore the estimates of total nitrogen deposition may be overestimated (Wetselaar
14 and Farquhar,, 1980; Bowden, 1986; Anderson and Levine, 1987; Schimel et al., 1988).
15 Melillo et al. (1989) indicate that losses of nitrogen from ecosystems in the form of N2O are
16 likely to be in the range of 2-4 kg N ha"1 yr"1. Higher levels of atmospheric nitrogen
17 deposition are also expected to lead to increased rates of N2O emissions.
18
August 1991
10-93
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
10.5 ECOSYSTEM EFFECTS-WETLANDS AND BOGS
10.5.1 Introduction
The diverse ecosystems that make up the biosphere interact through the cycling of
essential elements and compounds. The availability of these essential elements determines the
rates of biological processes within a given ecosystem. For example, the availability of
nitrogen in the form of NO3" (nitrate) or NH4+ (ammonium), which cycles through an
enormous atmospheric pool of N2, is an important determinant of the productivity of
ecosystems. Ecosystems interact and function in different ways with complex feedback
mechanisms; they influence the cycles of essential elements and, to some extent, even the
earth's climate.
Wetlands fulfill an important role in these global cycles as net sources and sinks for
biogenic gases. They transfer to the atmosphere globally significant quantities of CH4
(methane) (Harriss et al., 1982, 1985) and reduced sulfur gases (Steudler and Peterson,
1984). Elkins et al. (1978) discuss the possibility that coastal marshes may function as net
sinks for N2O (nitrous oxide). Because of the anaerobic nature of their waterlogged soils,
decomposition of organic matter in wetland soils is incomplete. Consequently, wetlands
function as sinks and long-term storage reservoirs for organic carbon. It has been estimated
that wetlands once sequestered a net of 57 to 83 X 106 metric tons of carbon per year
worldwide, although recent widespread drainage of wetland soils has shifted the carbon
balance (Armentano and Menges, 1986). Although this rate of carbon uptake is small in
comparison to other global carbon fluxes, such as the annual release of carbon from
combustion of fossil fuel (5-6 x 109 metric tons/y, Rotty, 1983) or the net uptake of CO2-C
by the ocean (1.6 X 109 metric tons/y, Tans et al., 1990), it is important when the net
balance between large fluxes is considered and it is certainly important over geologic time
scales (Armentano and Menges, 1986).
These gases (CH4, N2O and reduced sulfur compounds) modify atmospheric chemistry
and global climate. The destruction of ozone in the upper atmosphere by its reaction with
N2O is one example. Combustion sources are currently raising the atmospheric concentration
of N2O (Hao et al., 1987). The rise in anthropogenic releases of nitrogen oxides to the
atmosphere also increases the deposition of biologically available forms of nitrogen onto the ,
August 1991
10-94
DRAFT-DO NOT QUOTE OR CITE
-------
1 landscape with potential effects on productivity (or other aspects of function) and community
2 structure.
3 Locally, wetlands function as:
4 „ habitats for wildlife,
5 flood control systems,
6 stabilizers and sinks for sediments,
7 storage reservoirs for water, and
8 biological filters that maintain water quality.
9 Studies of riparian forests, for example, generally indicate that they exert a positive influence
10 on the water quality of receiving streams by intercepting and removing nutrients from runoff
11 (Yates and Sheridan, 1983; Brinson et al., 1984; Peterjohn and Correll, 1984; Quails, 1984).
12 And as sediment traps, salt marshes like those on the Louisiana coast can accumulate annually
13 an impressive 0.76 cm of sediment (DeLaune et al., 1983). These functions are a great
14 monetary value to society (Westman, 1977).
15 Wetlands also harbor a disproportionate (relative to habitat area) share of flora that are
16 threatened by extinction. Of the 130 plant species from the conterminous United States that
17 are formally listed as endangered or threatened (Code of Federal Regulations, 1987),
18 18 species (14%) occur principally in wetland habitats. On the national list of plant species
19 that are identified as endangered (Status LE or PE), threatened (Status LT or PT), or
20 potentially threatened (Status 1 or 2), 1776 species are listed for the conterminous United
21 States (Federal Register, 1985), and 302 of these (17%) occur principally in wetland habitats.
22 The national list of plant species that occur in wetlands includes 6,728 entries (Reed, 1988),
23 and since this list includes plant species found primarily in upland habitats as well as plants
24 from the entire United States and its territories, we can estimate conservatively that the
25 endangered, or potentially threatened wetland plant species represent an alarming 4.5%
26 (302/6,728) of this total.
27 Wetland plants are undoubtedly threatened because of loss of habitat, which in the
28 United States has been largely a consequence of agricultural development involving drainage
29 (Tiner, 1984). Total wetland area including intertidal and palustrine areas in the
30 conterminous United States (Figure 10-17) totaled 437,609 km2 during the mid-1950s and
31 decreased to 400,567 km2, or 5.1% of total land area, by the mid-1970s (Prayer et al.,
„ .. i •
32 1983). The net loss of wetland habitats during these two decades is equivalent to an annual
33 rate of loss of 1,852 km2/yr (=715 sq. miles per year). However, it can also be concluded
August 1991
10-95
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
that current rates of atmospheric nitrogen deposition in parts of Europe, elevated by
anthropogenic emissions, alter the competitive relationships among plants and threaten
wetland species adapted to infertile habitats. Those data are reviewed here, and on this basis
we can anticipate similar effects of atmospheric nitrogen deposition in the United States.
Figure 10-17. Map of the United States showing location of the major groups of inland
freshwater marshes (from Hofstetter, 1983, p. 213). Contours delineate
physiographic regions.
1 10.5.2 Atmospheric Nitrogen Inputs
2 Atmospheric nitrogen inputs occur as both wet and dry deposition. Most studies of
3 atmospheric nitrogen inputs into wetlands focus only on wet deposition or bulk deposition.
4 Accurate measurements of wet deposition are carried out by analyzing nitrogen in
5 precipitation immediately following a precipitation event. Frequently, however, rainfall is
6 accumulated over some period of time before it is analyzed, and the resulting measurement of
7 deposition rate is usually referred to as bulk deposition. Bulk deposition rates combine wet
8 deposition with some component of dry deposition. Where dry deposition has been carefully
August 1991
10-96
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
measured, it has been concluded that (1) the relative importance of wet and dry deposition
varies geographically, (2) that dry deposition can exceed wet deposition (Boring et al., 1988),
and (3) that bulk precipitation samplers underestimate the combined dry plus wet deposition
rate (Dillon et al., 1988). The available wet surface area of vegetation, onto which nitrogen
gases will diffuse, significantly affects the dry deposition rate (Heil et al., 1987). Levy and
Moxim (1987) modelled the fate of nitrogen oxide emissions to the atmosphere and concluded
that dry deposition accounts for greater than one-half of the total nitrogen oxide deposition in
North America.
The rate of bulk NO3" deposition has been shown to be positively correlated with the
concentration of nitrogen dioxide (NO2) in air. Press et al. (1986) measured atmospheric
concentrations of NQ2 and bulk deposition of NO3" at several sites in northern Britain for
18 months. NO2 concentrations (2 week averages) ranged from near zero to 25 Mg/m3 and
were correlated significantly (p< 0.001) with concentrations of NO3", collected in bulk
'samplers, that varied from near zero to about 3 mg N/litre.-
A third, and :rarely measured, mechanism of deposition that is locally important is the
interception or capture of fog or cloud droplets by vegetation. Lovett et al. (1982) estimated
that the cloud deposition of NO3" in an alpine habitat in New Hampshire was 101.5 kg
N ha"1 yr"1 compared to a bulk deposition rate of 23.4 kg N ha"1 yr"1. The same
phenomenon was observed by Woodin, and Lee (1987) who collected 1.45x as much water as
"throughflow" (collected beneath vegetation) passing through experimental Sphagnum mats in
the field as from adjacent bulk deposition gauges. Their data also suggests that the deposition
of solutes by this mechanism is important, and that bulk precipitation samplers underestimate
total deposition.
Table 10-17 summarizes several studies that report wet or bulk deposition rates of
nitrogen in North American wetlands, '' From- the data presented it may be concluded that bulk
deposition rates of NH4+, NO3", and organic nitrogen vary geographically and their relative
importance varies. In general, however, inputs of NO3", NH4+, and organic nitrogen are all
of the same order of magnitude, and their combined rate of deposition varies from 5.5 to
12.1 kg N ha"1 yr"1. Other studies, however, indicate that wet NO3" deposition alone
exceeds 15 kg N ha"1 yr"1 over most of the midwest, and 20 kg N ha"1 yr"1 in portions of the
northeast United States (Zemba et al., 1988).
August 1991
10-97
DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-17. BULK DEPOSITION OF NITROGEN IN NORTH AMERICAN
WETLANDS (kg N ha"1 yr'1)
Site
Chesapeake Bay, riverine tidal
emergent marsh
Massachusetts, salt marsh
Massachusetts, basin bog
Minnesota, spruce bog
Minnesota, spruce bog
Iowa, prairie marsh
Florida, everglades
Manitoba, emergent marsh
Ontario, poor fen
NH4+
2.7
1.4
2.5
1.7
3.0
4.0
3.0
NR
NR
NO3-
4.3
2.3
5.0
1.7,
2.0
4.0
9.6
NR
3.1
Org-N
4.7
3.9
NR
3.8
0.5
NR
NR
NR
NR
Tot-N
11.7
7.6
7.3
5.5
6.6-12.08
Reference
Jordan et al. (1983)
Valiela and Teal (1979)
Hemond(1983)
Verry and Timmpns (1982)
Urban and Eisenreich (1988)
Davis et al. (1983)
Flora and Rosendahl (1982)
Kadlec (1986)
Bayley et al. (1987)
NR — not reported.
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
Rates of nitrogen deposition, and NH4+ deposition in particular, in areas of Western
Europe are greater than in North America. In areas of Britain, bulk deposition rates of
43 and 46 kg ha"1 yr"1 have been reported (Press and Lee, 1982; Ferguson et al., 1984).
The combination of NO3" and NH4+ deposition down-wind of Manchester and Liverpool is
reported to be 32 kg N ha"1 yr'1 (Lee et al., 1986). Nitrogen deposition in fens near Utrecht
was 21 kg N ha"1 yr"1 of inorganic nitrogen and 3 to 5 kg N ha"1 yr"1 of organic nitrogen in
bulk precipitation and 18 kg N ha"1 yr"1 of inorganic nitrogen in dry deposition (Koerselman
et al., 1990). Roelofs (1983) reported that wet deposition alone of nitrogen in the
Netherlands averages 15 kg N ha"1 yr"1 and is as great as 20-60 kg N ha"1 yr"1 in areas of
intensive stockbreeding, 75-90% of this being deposited as NH4+. In Europe, 81 % of total
NH3 emissions are from livestock wastes, with the greatest emission densities concentrated in
The Netherlands and Belgium (Buijsman, 1987). Annual NH3 emissions from animal excreta
in The Netherlands are reported to be 230 kt yr"1 (Van der Molen et al., 1989) or about
60 kg ha"1 yr"1 country-wide. '
The chemistry of surface runoff from watersheds is probably of greater significance to
most wetlands than the chemistry of direct deposition, but the nitrogen load of surface runoff
probably increases with nitrogen deposition and with the size of the catchment area.
Atmospheric deposition accounts for a large fraction of the total nitrogen entering watersheds
(Robertson and Rosswall, 1986). Atmospheric deposition apparently has become a major
August 1991
•10-98
DRAFT-DO NOT QUOTE OR CITE
-------
1 source of NO3" to surface waters in North America, especially in the east and upper midwest
2 (Smith et al., 1987a), and increases in total nitrogen concentration at stream monitoring
I - ' •.
3 stations are strongly associated with high levels of atmospheric nitrate deposition (Smith
4 et al., 1987b). However, the direct contribution made by atmospheric deposition to the
5 nitrogen load in surface water is unknown. Measurements by Buell and Peters (1988) of
6 -stream chemistry in Georgia indicated that 93 % of the precipitation inputs of NH4+ and
7 NO3" were retained by the watershed. A study by Correll (1981) of mass nutrient balances
8 of a small watershed of the Rhode River estuary on the Chesapeake Bay showed that total wet
9 nitrogen deposition to 88 ha of tidal marshes and mudflats was 740 kg N (8.4 kg/ha) in
10 13 months compared to total nitrogen in runoff from 2,050 ha of watershed of 10,000 kg N.
11 Only about 7% (740 kg/10,740 kg) of the nitrogen entering the wetland was from direct
12 deposition. However, in as much as nitrogen deposition onto the watershed (8.4 kg/ha x
13 2,050 ha = 17,220 kg) exceeded, total runoff from the watershed to the wetland (10,000 kg),
. - -f • - '-•-'- •- •
14 deposition could have contributed indirectly through runoff the majority of nitrogen entering
15 the wetland. But the contributions of other nitrogen sources to runoff, such as fixation,
16 fertilizer, and animal waste, were not given. . ,
17 ' t , . , ;
18 10.5.3 The Wetland Nitrogen Cycle
19 The feature of wetlands that sets them apart from terrestrial ecosystems is the anaerobic
20 (oxygen-free) nature of their waterlogged soils which alters the relative importance of various
21 microbial transformations of inorganic and organic nitrogen compounds. Generally, the
22 ; absence of O2 retards the decomposition of organic matter (Tate, 1979; DeLaune et al.,
23 1981; van der Valk and Attiwill, 1983; Godshalk and Wetzel, 1978; Clark and Gilmour,
24 1983). Complex aromatic ring structures are more resistant to microbial attack under anoxic
25 conditions (Tate, 1979), leading to the formation and buildup of peat in wetland
26 environments. Anoxic soils also favor the rapid conversion of N03~ to N2Q (nitrous oxide)
27 or N2 (nitrogen gas). This process is accomplished by bacteria and is referred to as
28 - denitrification or dissimilatory nitrate reduction, and it results in quantitatively important
29 losses of nitrogen from wetland ecosystems. Finally, the hydrology of wetlands favors
30 • diffusive exchanges of nitrogen compounds to and from sediments and advective transport
31 (carried by water) of nitrogen compounds between ecosystems. This often results in
August 1991
10-99
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
movements of NH4+ from anoxic sediments to the oxidized surface sediment or water
column, where nitrification (the oxidation of NH4+ to NO3" by bacteria) can occur, and the
return movement of NO3" to the anoxic sediment layers where denitrification can occur. The
nitrogen cycle in wetlands has been reviewed recently by Reddy and Patrick (1984), Savant
and De Datta (1982), and Bowden (1987). Important steps in the nitrogen cycle are
summarized in Section 10.1.3.
Table 10-18 the nitrogen budgets are presented of wetlands that exhibit a wide range of
nitrogen inputs. The two bog sites (Table 10-18) are representative of wetlands that contain
plant species that are adapted to low levels of nitrogen. They are examples of ombrotrophic
bogs, meaning that they receive nutrients exclusively from precipitation. They develop where
precipitation exceeds evapotranspiration and where there is some impediment to drainage of
the surplus water (Mitsch and Gosselink, 1986). Bogs are dominated by Sphagnum spp. and
may be sparsely forested. The Sphagnum builds a dense layer of peat, creating a surface that
contains no mineral sediment. The peat in ombrotrophic bogs is raised above the elevation of
the surrounding land so that they receive neither runoff from uplands nor inputs from ground
water. Peat forming bog ecosystems are widely distributed throughout the northern
hemisphere, but they are most common in formerly glaciated regions. The distribution of
peafland area in North America is shown in Figure 10-18. The bog ecosystems represented
in Table 10-18 are located in Minnesota (Urban and Eisenreich, 1988) and Massachusetts
(Hemond, 1983).
In bog ecosystems, the most important nitrogen inputs are from wet and dry deposition
(see the line labelled precipitation in Table 10-18). The total input of nitrogen in these
examples is about 10 kg N ha"1 yr"1, and atmospheric deposition accounts for most of this
(Urban and Eisenreich, 1988; Hemond, 1983). Also note that the total nitrogen outputs from
the system are approximately 4 kg N ha"1 yr"1. The outputs are accounted for by
denitrification (1 to 1.8 kg N ha"1 yr"1) and by export in runoff of dissolved inorganic
nitrogen (as NH4+) and dissolved organic nitrogen (DON). No export of particulate organic
nitrogen was reported; nitrogen accumulated in plant tissues is largely recycled within the
bog.
Bog wetlands are representative of one end of a continuum, but there are also other
wetlands where atmospheric nitrogen deposition represents a significant fraction of the totaf
August 1991
10-100
DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-18 NITROGEN BUDGETS OF SELECTED WETLANDS
(kg N ha'1 yr'1)
Location and Wetland Type:
INPUTS:
Precipitation
Fixation
Surf. , ground or tidal water
Total
INTERNAL CYCLE:
Plant assimilation
Mineralization
OUTPUTS:
Denitrification
Ammonia volatilization
Surf, or subsurface DIN export
Surf, or subsurface ON export
Total
UK
Salt
Marsh1
NR
3.36
43.4*
225.4
194.9
3.78
NR
2.4®
43.0®
MA
Salt
Marsh2
7.9
68.0
668.0
743.9
214.0s
193. 0$
143.0
0.35
102.0
552.0
797.4
Dutch
Rech.
Fen3
43. 71
2.1
7.3
53.1
274.0f
244.0*
1.4
NR
2.1
45. 8§
49.3
Dutch
Disc.
Fen3
42.01
12.7
20.9
75.6
90.0t
79.0*
1.1
NR
6.7
80.4§
88.2
French
Heath4
8.1
1.3
0
9.4
82.0
74.0
NR
NR
3.0
3.0
MA
Bog5
7.5
3.36
0
1.09
38.0
26.0
1.0
Trace
2.0
1.0
4.0
MN
Bog6
8.6
0.5
0
0.91
66.0
50.0
1.8
NR
0
2.0
3.8
DIN = Dissolved inorganic nitrogen.
ON = Dissolved and particulate organic nitrogen.
NR = Not Reported.
'Abd. Aziz and Nedwell (1986a,b): salt marsh dominated by Puccinellia maritima (a grass).
2Valiela and Teal (1979): salt marsh dominated by Spartina alterniflora.
3Koerselman et al. (1990): Dutch eutrophic recharge and mesotrophic discharge fens, respectively.
4Roze (1988): mesophilous heathland (shrub bog) dominated by Erica ciliaris (heath) and Ulex minor.
5Urban and Eisenreich (1988): ombrotrophic Sphagnum bog forested with black spruce (Picea mariand) and
with an understory of shrubs and sedges.
6Hemond (1983): ombrotrophic bog dominated by Sphagnum.
Calculated from Morris et al. (1984) and Valiela et al. (1984).
^Represents the net exchange of NO3- (the major component) and small particulate organic nitrogen rather than
an absolute rate.
®Represents the net exchange of DON (the major component), NH4+, and large particulate organic nitrogen
rather than an absolute rate.
^Includes bulk plus dry deposition of inorganic and organic nitrogen.
tFrom Verhoeven et al. (1988), assuming a root:shoot quotient of 1.0.
*From Verhoeven et al. (1988).
§Includes primarily hay harvested by mowing.
1 input of inorganic nitrogen. For example, wetfall contributed more than 95% of the NH4+
2 and NO3" entering the 1,000 km2 Shark River Slough, the major fresh water drainage of
3 Everglades National Park (Flora and Rosendahl, 1982). However, the importance of organic
4 nitrogen in the surface inflow may be considerable, depending on how easily or rapidly it is
August 1991
10-101
DRAFT-DO NOT QUOTE OR CITE
-------
O.S - 10%
P««tJ*nd Area
PtttUnd
Figure 10-18. Distribution of North American peatlands (from Mitsch and Gosselink.
1986, p. 288).
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
mineralized by the microbial community. In this ecosystem, rainfall is about 84% of total
water input, and one can generalize that the significance of atmospheric nitrogen deposition
increases in wetlands as rainfall increases as a fraction of the total water budget.
The French heathland or shrub bog (Table 10-18) is another example of a wetland with
low nitrogen inputs and outputs, but with an intermediate rate of internal cycling. The
moderate size of the internal nitrogen cycle depends on the accumulation of a large quantity
of organic nitrogen in the soil humus (Roze, 1988). A fraction of this organic pool
mineralizes each year and is assimilated by the plant community. Organic and inorganic
nitrogen in the soil is about 91% of total nitrogen in this heathland ecosystem, with the
remaining 9% being contained within the plant biomass. A moderate rate of nitrogen
mineralization in the soil is balanced by assimilation by the plant community, and nitrogen is
largely conserved within the ecosystem.
In the Dutch fens (Table 10-18), the inputs and outputs of nitrogen are intermediate
between those of the bogs and salt marshes. Both fens are influenced by their close
proximity to heavily fertilized pastures, by atmospheric nitrogen deposition, and by annual
mowing and harvest of aboveground vegetation. The fen that occupies a site of ground water
August 1991
10-102 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
recharge is influenced by water that is diverted from the highly polluted river Vecht during
periods of high evapotranspiration, and the discharge fen is influenced by nutrients in ground
water (Verhoeven et al., 1988). However, atmospheric nitrogen deposition in these fens
supplies more nitrogen than all other inputs combined (Table 10-18).
The coastal salt marsh ecosystems in Table 10-18 are characteristic of wetlands that are
adapted to large nitrogen inputs. Coastal salt marshes have a temperate, worldwide
distribution. They exist within the intertidal zone and are alternately flooded and drained
daily by the action of the tides. The example from Massachusetts is a salt marsh dominated
by the grass Spartina alterniflora (Valiela and Teal, 1979). The salt marsh example from the
United Kingdom in Essex is dominated by the grass Puccinellia maritima (Abd. Aziz and
Nedwell, 1986b).
In salt marsh ecosystems, the most important nitrogen inputs are from those brought
into the marsh in tidal water and, in some case's, ground water. Input of pafticulate organic
nitrogen from sedimentation and/or NO3" is apparently the dominant mechanism by which
these ecosystems remove nitrogen from surface water, since the diffusion gradients for NH4+
and DON normally favor diffusion out of the sediment. These surface and ground water
sources of nitrogen are one to two orders of magnitude greater than inputs from precipitation
(Table 10-18). In the Massachusetts salt marsh, ground water inputs of NO3" and DON are
important and account for 60 and 56 kg N ha"1 yr"1, respectively, of the total inputs (Valiela
and Teal, 1979). In contrast, the Essex (UK) marsh is not influenced by ground water (Abd.
Aziz and Nedwell, 1986b). Both salt marshes have large nitrogen inputs from tidal water,
and in the Massachusetts marsh these are largely as NH4+ (54 kg N ha"1 yr"1), DON
(337 kg N ha"1 yr"1), and particulate organic nitrogen (139 kg N ha"1 yr"1) (Valiela and Teal,
1979). There are additional inputs and outputs, such as deposition of bird faeces and
shellfish harvest, but these are insignificant in comparison to other rates (Valiela and Teal,
1979).
The large inputs of nitrogen in salt marshes are balanced by equally large outputs
(Table 10-18), but there are important transformations that take place within the marsh.
Denitrification accounts for 17.9% of the total nitrogen loss from the Massachusetts marsh.
Because the denitrification rate is greater than the combined inputs of NO3"' this implies that
rates of nitrification are large. In both marshes, the greatest nitrogen losses occur in tidal
August 1991
10-103
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
S
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
water exchange, and in the Massachusetts marsh there is a net loss of all forms of dissolved
nitrogen in tidal water. The Massachusetts marsh exports large amounts of NH4+ (73 kg
N ha"1 yr"1), NO3" (25 kg N ha"1 yr'1), DON (380 kg N ha"1 yr"1), and particulate organic
nitrogen (17 kg N ha"1 yr"1) (Valiela and Teal, 1979).
Nitrogen inputs and outputs in tidal water were given as net exchanges of different
nitrogen components in the Essex (UK) study (Abd. Aziz and Nedwell, 1986b) rather than
absolute rates. This is the reason for the discrepancy in the rates of tidal water imports and
exports of nitrogen in the Essex and Massachusetts marshes (Table 10-18). However, valid
comparisons can be made of the net exchanges. There is a large net export of DON
(43 kg N ha"1 yr"1) from the Essex marsh (Abd. Aziz and Nedwell, 1986b), and this is
consistent with the net DON loss in tidal water of 45 kg N ha"1 yr"1 from the Massachusetts
marsh (Valiela and Teal, 1979). The marshes differ in the net tidal water exchanges of other
forms of nitrogen.
The rate of internal nitrogen cycling (assimilation and mineralization) within ecosystems
is directly proportional to the rate of primary production (e.g., Verhoeven and Arts, 1987),
although high rates of productivity can be supported by high external nutrient inputs when
conditions are unfavorable for high mineralization rates (Verhoeven et al., 1988).
Mineralization rates differ greatly between the wetland types represented in Table 10-18.
Nitrogen assimilation by the plant communities varies from 38 to 66 kg N ha"1 yr"1 in the
bog ecosystems compared to 225 to 274 kg N ha"1 yr"1 in the salt marsh and fen ecosystems,
respectively. The nitrogen cycle in the bog and heathland ecosystems is largely closed
(Figure 10-19). In contrast, the nitrogen cycle in salt marshes and fens is open, and there is
a great exchange of nitrogen with adjacent systems (Figure 10-19). In all these ecosystems,
the rate of nitrogen mineralization almost balances plant assimilation in the manner of a
closed cycle (Table 10-18). However, it is unlikely that the salt marsh could function as a
closed system and maintain its productivity or community structure. Likewise, it is unlikely
that the bog ecosystem could maintain its community structure if the nitrogen inputs were
greatly increased by some means. In general, as the input rate of nitrogen increases there are
concomitant increases in the output rate and magnitude of the internal cycle (Table 10-18).
August 1991
10-104
DRAFT-DO NOT QUOTE OR CITE
-------
Low N-inputs
Low N-outputs
Oligolrophic habitats
eg. ombrotrophic bogs
Eutrophic habitats
eg. salt marshes
Assimilation Mineralization
' Low
' Productivity
Internal Cycling
Moderate
Species Diversity
Low N-inputs
Assimilation
High N-inputs
Low N-outputs
•K
Moderate
Productivity
Internal Cycling
High
Mineralization Species Diversity
High N-outputs
Low
Species Diversity
High
Species Diversity
Internal Cycling
Assimilation
Mineralization
Figure 10-19. Conceptual relationships among trends in nitrogen cycling, productivity,
and species diversity along a gradient from oligotrophic (nutrient-poor) to
eutrophic (nutrient-rich) habitats.
1
2
3
4
5
6
In ecosystems with closed nutrient cycles and small rates of internal cycling, like bogs, if
nitrogen loadings increase significantly, then we can predict that productivity will increase,
but as will be discussed later, the increased productivity will be accompanied by changes in
species composition to those adapted to an elevated nutrient regime (Figure 10-19).
August 1991
10-105
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
10.5.4 Effects of Nitrogen Loading on Wetland Plant Communities
10.5.4.1 Effects on Primary Production
Numerous field experiments involving nitrogen fertilization have documented that
primary production in wetland ecosystems is commonly limited by the availability of
nitrogen. Results of this type of experiment are presented in Table 10-19. In all of the
fertilization experiments included in the table, only sewage sludge, urea, or mineral nitrogen
in the form of NH4+ or NO3~ were applied. Except in the case of sewage sludge
applications, where the numerous elements contained in sludge preclude attributing the results
to any specific element, the stimulation of growth that was observed can be attributed solely
to application of nitrogen. Rates of application ranged from 7 to 3,120 kg N ha"1 yr"1
(Table 10-19), and in most studies these have been 1 to 2 orders of magnitude greater than
rates of atmospheric deposition (Table 10-17). These applications stimulated increases in
standing biomass by 6 to 413% (Table 10-19).
Several studies have investigated the effects of different nitrogen sources. Cargill and
Jefferies (1984) found that applications of NH4+ increased production of Puccinellia
phiyganodes (a grass) in a sub-arctic salt marsh by 175%, while equivalent applications of
NO3" increased production by only 73%. Applications of NO3" were perhaps less effective
than NH4+ because of denitrification of NO3" by bacteria in the anaerobic marsh sediments.
This demonstrates the importance of competition between plants and microbes for specific
inorganic nitrogen compounds, with plants being the best competitors for NH4+.
The greatest stimulation of growth is often achieved when nitrogen applications are
combined with applications of other nutrients. In the study of Cargill and Jefferies (1984)
applications of Pf (inorganic phosphate) combined with NH4+ stimulated production to a
greater extent than NH4+ alone. Sanville (1988) observed that combinations of nitrogen, in
the form of urea, and Pj stimulated production in a Sphagnum bog to a greater extent than
nitrogen applications alone, and that singular additions of PJJ had no significant effect on
growth. These results demonstrate that other nutrients, PA in these examples, become
secondarily limiting after nitrogen applications reach a threshold.
In one study of a wet heathland in the central Netherlands, total aboveground biomass
failed to respond on experimental sites fertilized for 3 yr at a rate of 200 kg N ha"1 yr"1, but
sites fertilized with 40 kg P ha"1 yr"1 did show a significant increase in biomass (Aerts and
August 1991
10-106
DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-19.
Salt Marsh Ecosystems
Spartina
Spartina
Spartina
Spartina
Spartina
Spartina
Spartina
Pucdnellia
Puccinellia
Car ex.
Panicum hemitomon
Panicum hemitomon
Typha glauca
Sparganium eurycarpum
bog
bog
fen .,
wet grassland
RESULTS OF NITROGEN FERTILIZATION EXPERIMENTS
IN WETLAND ECOSYSTEMS
Rate of N
Application
(kg ha'1 yr4)
200
200
220 ...
650
670
1,040
3,120
320
320
320
30
100
1,350
1,350
300
7
450
450
Length of
Study (yr)
1
1
3
3
2
1
2
2
2
2
1
1
2
2
1
1
1
1
Control1
Biomass
(g/m2)
1,660
816
320
320
250
450
235
64
64
65
1,320
1,320
1,726
637
180
'200
350
400
Percent?
Increase
16
25
131
269
120
100
413
175
73
146
6
42
36
86
25
10
57
68
N-form
Applied
NH4+
NH4N03
Sludge
Urea
Sludge
NH4+
NH4+
NH/'
N03-
NH4+
NH4+
NH4+
NH4NO3
NH4N03
Urea
' Sludge
Mineral-N
Mineral-N
Reference
Patrick and Delaune (1976)
Gallagher (1975)
Valiela et al. (1975)
Valiela et al. (1975)
Valiela and Teal (1974)
Haines (1979)
Morris (1988)
Cargill and Jefferies (1984)
Cargill and Jefferies (1984)
Cargill and Jefferies (1984)
DeLaune et al. (1986)
DeLaune et al. (1986)
Neely and Davis (1985a)
Neely and Davis (1985a)
Sanville (1988)
Sanville (1988)
Vermeer (1986)
Vermeer (1986)
'Control biomass is the .maximum, nonfertilized aboveground standing crop
2Percent increase indicates the response of control biomass during the year of fertilization at the indicated rate of
application, computed as lOOx(Experimental-Control)/Control
1 Berendse, 1988). Thus, wetlands are not universally limited by nitrogen. However, as
i • , '
2 discussed above (Atmospheric Nitrogen Inputs - Section 10.5.2) The Netherlands is an area
3 of extreme high nitrogen deposition, and the threshold for nitrogen limitation is perhaps
4 exceeded by anthropogenic inputs in this area.
5 Fertilization experiments of salt marshes in Massachusetts by Valiela and Teal (1974)
6 and in Louisiana by Patrick and Delaune (1976) involving singular applications of either
7 nitrogen or Pj demonstrated that primary production was stimulated by nitrogen and not by
8 phosphorus. Vermeer (1986) obtained the same result in freshwater fen and wet grassland
9 communities in The Netherlands. However, fertilization with nitrogen increased the biomass
10 and dominance of grasses at the expense of other species in fen and wet grassland
11 communities. Some Equisetum spp. (horesetail) had a smaller biomass contribution upon
August 1991
10-107
DRAFT-DO NOT QUOTE OR CITE
-------
1 fertilization. This tendency toward a change in species composition or dominance has been
2 observed in other fertilization experiments also. Jefferies and Perkins (1977) found species-
3 specific changes in stem density at a Norfolk, England salt marsh after fertilizing monthly
4 with 610 kg NO3"-N ha"1 yr"1 or 680 kg NH4+-N ha"1 yr"1 over a period of 3-4 years.
5 A final conclusion of the data in Table 10-19 is that the stimulation of primary
6 production by nitrogen applications is not a linear function of the rate of nitrogen application.
7 This can be seen by comparing the results of fertilization studies of Spartina (Table 10-19).
8 The greatest increase in standing biomass, both in terms of absolute amount and in terms of
9 the percent increase, was obtained in studies where the control biomass was low. This
10 implies that the in situ nitrogen supply in some wetlands already is near a threshold where
11 other factors become limiting. Ultimately, available light energy, water, and temperature are
12 the limiting factors.
13 The data included in Table 10-19 pertain to growth of aboveground biomass only. In
14 several of these studies, measurements of belowground biomass were also made (Valiela and
15 Teal, 1974; Haines, 1979; Valiela et al., 1976; Gallagher, 1975). Results were variable with
16 some studies showing a small decrease in living belowground biomass (Valiela et al., 1976)
17 and others showing small increases in belowground macroorganic matter (Gallagher, 1975),
18 or no change (Valiela and Teal, 1974). The normal technique of coring sediments to measure.
19 belowground production is subject to great error (Singh et al., 1984). However, the evidence
20 from controlled growth experiments (Morris, 1982; Steen, 1984) is clear that the response of
21 leaf growth to increased nitrogen supply is much greater than the response of roots.
22 It should be emphasized that all of the fertilization studies summarized in Table 10-19
23 are short term results in which nitrogen was applied for 3 years or less. We cannot assume
24 that long-term nitrogen applications will yield the same results. Studies of several wetland
25 ecosystems that have been fertilized for long periods by increased atmospheric inputs indicate
26 that changes in species composition and succession accompany the increases in nitrogen
27 loadings and primary production. These studies are summarized below.
28 One implication of a long-term increase in leaf growth is that the demand for mineral
29 elements and water from the soil will increase. Howes et al. (1986) observed that the rate of
30 evapotranspiration increased from a salt marsh dominated by Spartina alterniflora in sites
31 where aboveground biomass was increased by nitrogen fertilization. Increased
August 1991
10-108
DRAFT-DO NOT QUOTE OR CITE
-------
1 evapotranspiration can influence the direction of succession of some wetlands by altering the
2 water balance of the soil. The feasibility of this mechanism to alter bog succession was
3 demonstrated in a model by Logofet and Alexandrov (1984). Their model suggests that
4 nitrogen inputs greater than a threshold of 7 kg N ha'1 yr"1 can change the direction of
5 succession from that of an open oligotrophic bog to a mesotrophic bog dominated by trees.
6 Furthermore, in flowing water systems, like salt marshes, an increase in aboveground
7 production should lead to an increased export from the system of'nutrients that are
8 incorporated in or leached from aboveground biomass. Therefore, the long-term ecosystem
9 and community responses to increased inputs of nitrogen can not be predicted from results of
10 short term field experiments like those summarized in Table 10-19.
11
12 10.5.4.2 The Fate of Added Mineral Nitrogen
13 Experiments in the field and laboratory have followed the fate of applied nitrogen by
14 using 15N as a tracer. 15N is a stable isotope comprising 0.37% of naturally occurring
15 nitrogen. It can be quantified together with the more common isotope of nitrogen, 14N, with
16 a mass spectrometer and is used experimentally much like radioactive isotopes except that
17 15N is normally used in greater than trace amounts due to the lower sensitivity of the
18 instrumentation used to detect it.
19 Experiments in which different mineral forms of 15N were added to sediments in the
20 absence of plants demonstrate that mineral nitrogen is rapidly used by the microbial
21 community. Smith and Delaune (1985) added the equivalent of 100 kg N ha"1 in one
22 application as 15NH4+ to sediments of a shallow saline lake. They found 15 days after the
23 addition, 20% had been converted to organic nitrogen in the sediment, and the fraction in
24 organic matter remained constant at this level for the remaining 337 days of the experiment.
25 The amount of 15NH4+ in the sediment decreased exponentially to a nondetectable level by
26 Day 200. Diffusion of NH4+ into the water column and denitrification accounted for a loss
27 of 80% of the 15NH4+ from the sediment.
28 Lindau et al. (1988) made single additions of either 15NO3" or 15NH4+, equivalent to
29 100 kg N ha"1, to the floodwater within chambers containing swamp sediment. By Day 27,
30 only 39.6% and 6.2% of the 15N from NH4+ and NO3", respectively, remained in the
31 sediment and overlying water column. The remaining fractions had been lost from the
August 1991
10-109 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
chambers by denitrification. The loss of 60% of the applied 15NH4+ within 27 days
demonstrates that NH4+ can be rapidly converted to NO3" by nitrifying bacteria in aerobic
parts of the system, and that NO3" diffuses into the anaerobic sediments where denitrification
occurs. Nitrification was apparently the rate limiting step since the loss of 15N by
denitrification was more rapid when it was applied as NO3".
DeBusk and Reddy (1987) made single additions of 15NH4+ to the floodwater above
cores of sediments taken from swamps that had been receiving primary wastewater effluent
for 2 and 50 years prior to the experiment. The rate of application was equivalent to 15 kg
N ha"2. After 21 days, 0.5 to 2.3% of the added nitrogen was recovered in the flood water,
largely as NO3", and 13.6 to 17.8% in the sediment, largely as organic matter. The
remaining 80% was apparently lost by denitrification, indicating that conversion of NH4+ to
NO3" and diffusion of NO3" to anaerobic sites of denitrification is rapid. This result is
consistent with that of Lindau et al. (1988). Furthermore, there was no difference in the
response of the two sediment types, which demonstrates that the nitrification-denitrificatibn
potential of sediments is unchanged in sediment receiving sewage effluent for 50 years.
However, the bacteria in the sediments must have a continuous supply of suitable carbon
substrates as well as nitrogen to sustain continuous nitrification-denitrification reactions.
Short term measurements of slurrys of marl and peat sediments from the Florida
Everglades (Gordon et al., 1986) demonstrated that 10-34% of NO3" added at levels of
10 and 100 #M (1 iM. = 14 #g N/litre) was rapidly denitrified within 24 h. Denitrification
rates decreased following this initial burst of activity as the balance of the added NO3" was
converted to NH4+. This experiment suggests' that the process of dissimilatory nitrate
reduction to ammonium (reammonification) competes successfully with the denitrification
process. However, this experiment was conducted on sediment slurrys that were incubated
under a nitrogen atmosphere which prevented nitrification reactions from occurring. Under
an oxygen atmosphere, nitrification would have generated a continuous supply of NO3" and
denitrification would then have consumed a greater fraction of the NO3" over time.
The behavior of mineral nitrogen applied to vegetated wetland sediments is quite
different from the results described above and indicates that plants successfully compete with
microbes for mineral nitrogen. Delaune et al. (1983) followed the fate of 15NH4+ placed
below the soil surface in a Louisiana salt marsh dominated by Spartina alterniflora. The
August 1991
10-110 DRAFT-DO NOT QUOTE OR CITE
-------
1 singular application of 15NH4+ was equivalent to 72 kg N ha"1. At the end of .the first
2 growing season, 93% of the added nitrogen was recovered in aboveground biomass, roots,
3 and soil. An average of 28% was in aboveground biomass and 65% was in soil and
4 belowground biomass. The high rate of recovery of 15N in vegetation and soil is consistent
5 with results of Buresh et al. (1981) and Patrick and Delaune (1976). In the study of Delaune
6 et al. (1983), 15N recovered in soil and belowground biomass declined to 50% by the end of
7 the second growing season and to 43% by the end of the third growing season. Nitrogen in
8 aboveground biomass decreased to 1.2% of original 15N by the end of the third growing
9 season. The annual declines were postulated to have occurred due to the loss of nitrogen
10 from the leaves, either by physical transport of aboveground plant material off the site or by
11 decomposition of leaf material at the sediment surface followed by nitrification-denitrification
12 reactions. Similar results were obtained in a freshwater marsh dominated by Panicum
13 hemitomon (maiden cane). Delaune et al. (1986) added 30 kg ha'1 of 15NH4+-N to
14 sediments and recovered a mean of 80% in the combined aboveground (18%) and
15 belowground biomass and soil (62%) at the end of the first growing season. ,
16 Dean and Biesboer (1985) applied 15NH4+ to the flood water in cylinders containing
17 sediment only and in cylinders containing Typha latifolia (broadleaved cattail). Additions
18 were made biweekly during a single growing season for a total application equivalent to
19 82 kg N ha"1 season"1., At the end of the growing season, 3 weeks after the last addition^
20 75.3% of added 15N was recovered in the plant-soil system. A total of 53.6% was contained
21 in the plants, including both above and belowground biomass, while 21.7%( was contained in
22 the soil. In the sediment-only system, only 34.6% of the added 15N was recovered; most of
23 this, 33% of the added 15N, was in the sediment. The remaining 65.4% was thought to have
24 been lost through nitrification-denitrification reactions.
25 The experiments discussed above indicate that plant biomass is the major sink for free
26 NH4+, and that in the absence of plants, the major fate is nitrification-denitrification. It
27 should be emphasized that the nitrification-denitrification process can dominate only in .
28 environments, like wetlands, that have separate and distinct aerobic and anoxic zones of
29 microbial activity where solutes freely, diffuse between them.
30
31 .-.'..-
August 1991
,10-111 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
10.5.4.3 Effects of Nitrogen Loading on Microbial Processes
Changes in deposition rate and the chemical form of nitrogen in deposition can
potentially influence microbial processes and details of the internal nitrogen cycle of
wetlands. For instance, decomposition rate is sensitive to the nitrogen concentration of
decomposing tissues and of the surrounding environment. Tissues with elevated nitrogen
concentrations normally are observed to decompose at a faster rate than tissues containing low
nitrogen concentrations (Marinucci et al., 1983; Neely and Davis, 1985b). The difference in
decomposition rates can be impressive. For example, litter fr6m N-fertilized Spartina
alterniflom decomposed 50% faster than control litter (Marinucci et al., 1983).
The dynamics of nitrogen within decomposing litter is also sensitive to the litter's
nitrogen status. That is, litter of low original nitrogen content often acts as a net nitrogen
sink during the first months of decomposition, whereas nitrogen-rich litter is likely to be a
nutrient exporter rather than an accumulator during decomposition (Neely and Davis, 1985b).
; E ' *' ' .
There is some controversy about the mechanism of nitrogen immobilization (Bosatta and
Staaf, 1982; Aber and Melillo, 1982; Bosatta and Berendse, 1984), but its importance to the
wetland nitrogen cycle is recognized (Brinson, 1977; Morris and Lajtha, 1986; Damman,
1988).
Microbial nitrogen transformations are also affected by the nitrogen status of the
environment. It is well known that NH4+ inhibits the activity of nitrogen fixing bacteria
(diazotrophs) (Buresh et al., 1980). It is thought that NH4+ represses synthesis by bacteria
of the nitrogenase enzyme (the enzyme in bacteria that accomplishes the transformation).
There may be direct inhibition by NH4+ of enzyme activity as suggested by Yoch and
Whiting (1986). Kolb and Martin (1988) observed a decrease in nitrogenase activity as well
as the proportion of diazotrophs among the heterotrophic bacteria in soil after application of
NH4NO3. They suggested that the decrease in proportion of diazotrophs represents a
competitive suppression by non-diazotrophs in the presence of combined nitrogen (NH4+ or
NO3~). bicker and Smith (1980) observed a similar repression of nitrogen fixation in salt
marsh sediments amended with either NH4+ or NO3".
Acidification, which may'be caused by deposition of NOX or NH4+, can impact the
nitrogen cycle. Decomposition rate is decreased by acidification (Leuven and Wolfs, 1988;
Hendrickson, 1985), but the degree of inhibition is dependent upon the buffering capacity of
August 1991
10-112 DRAFT-DO NOT QUOTE OR CITE
-------
1 the litter (Gallagher etal., 1987). Nitrification is also affected by acidification. Nitrification
2 was inhibited at pH 4-5 in cypress swamps (Dierberg and Brezonik, 1982), and at pH 5.4 to
3 5.7 in lakes (Rudd et al., 1988). Acidification blocks the nitrogen cycle by inhibiting
4 nitrification and leads to an accumulation of NH4+ (Roelofs, 1986; Schuurkes et al., 1986,
5 1987; Rudd et al., 1988). Also, the ratio of N2O:N2 produced by denitrifying bacteria is
6 apparently pH sensitive with little N2O being produced under anoxic conditions at pH 7 and
7 almost 100% at pH 5 (Focht, 1974). This is significant, because a shift to N2O production
8 upon acidification of the environment could have a deleterious effect on stratospheric ozone.
9 Finally, NO3" and NH4+ have been shown to influence the relative and absolute ,
10 production of endproducts of dissimilatory nitrate reduction (Blackmer and Bremner, 1978;
11 Knowles, 1982; Prakasam and Krup, 1982). King and Nedwell (1985) observed
12 approximately equal reduction to either NH4+ or N2O (in the presence of acetylene the gas
13 added to assay the rate of production of N2O) in sediment slurrys incubated anaerobically
14 with 250 /iM NO3~. As the nitrate concentration was increased, up to 2 mM (1 mM =
15 14 mg N/litre), the proportion of the nitrate which was denitrified to N2O increased up to
16 83%. High nitrate concentrations have also been shown to favor N2O production and inhibit
17 N2 production, perhaps due to the competitive role that exists between NO3~ and N2O
18 terminal electron acceptors during anaerobic respiration (Cho and Sakdinan, 1978; Blackmer
19 and Bremner, 1978). Seitzinger et al. (1983 and 1984) observed higher ratios of N2O:N2
20 production and higher absolute rates of N2O production from eutrophic sediments than from
21 unpolluted sediments of Narragansett Bay, RI. Smith and Delaune (1983) reported that N2O
22 production from salt marsh and brackish marsh soils increased from 0.22 and 0.04 mg
23 N2O-N m"2 day"1, respectively, to 1.5 and 2.9 mg N2O-N m"2 day"1 after amending the
24 sediments with 1.2-1.5 g NH4+-N m"2. Others (Betlach and Tiedje, 1981), however, failed
25 to observe an inhibition of N2O reduction in the presence of NO3". Little is known about the
' ' 4 | t '*
26 significance of this process in general or the potential for NO3~ or NH4 in deposition to
27 alter natural rates of N2O production. Only a small fraction of depositional nitrogen inputs
28 are likely to be evolved as N2O. For example, Pedrazzini and Moore (1983) recovered only
29 0.39% of fertilizer-N as N2O from submerged soils amended with 34 g NO3"-N m"2 and 12 g
30 NH4+-N m"2 in the laboratory. However, on a global basis, even small changes in the
August 1991
10-113
DRAFT-DO NOT QUOTE OR CITE
-------
1 production of N2O are potentially significant considering the role of N2O in the destruction
2 of stratospheric O3 (Crutzen, 1970; Hahn and Crutzen, 1982).
3
4 10.5.4.4 Effects on Biotic Diversity and Ecosystem Structure
5 In the introduction it was pointed out that wetlands harbor about 17% of the total
6 number of plant species formally listed as endangered in the United States. While it is
7 beyond the scope of this review to survey the physiological ecology of these wetland plants,
8 several species on this list are widely recognized to be adapted to nitrogen-poor or infertile
9 environments. These include the isoetids (Boston, 1986) and the insectivorous plants (Keddy
10 and Wisheu, 1989; Moore et al., 1989; Wisheu and Keddy, 1989), like the endangered green
11 pitcher plant, Sarracenis oreophila. In eastern Canadian wetlands, nationally rare species are
12 found principally on infertile sites (Moore et al., 1989, Wisheu and Keddy, 1989).
13 Therefore, management practices should recognize that alterations in competitive relationships
14 between species occur when the fertility of the environment changes.
15 These assertions are supported by research on floristic changes related to nitrogen
16 deposition in central Europe. Ellenberg (1988) surveyed the nitrogen requirements of
17 1,805 plant species from West Germany and concluded that 50% can compete successfully
18 only in habitats that are deficient in nitrogen supply. Furthermore, of the threatened plants,
19 75 to 80% are indicator species for habitats of poor nitrogen supply. When stratified by
20 ecosystem type, it is also clear that the trend of rare species occurring with greater frequency
21 in nitrogen-poor habitats is a common phenomenon across many ecosystem types
22 (Figure 10-20, 10-21).
23 There is a history in western Europe of changes in wetland community composition that
24 are thought to result from deposition of atmospheric pollutants. Sphagnum species are largely
25 absent from ombrotrophic peat bogs in areas of Britain where they were once common
26 (Tallis, 1964; Ferguson and Lee, 1980; Ferguson et al., 1984; Lee et al., 1986).
27 Ombrotrophic wetlands downwind of the Manchester and Liverpool conurbations have been
28 extensively modified by atmospheric pollution for greater than 200 yr, with the virtual
29 elimination of the dominant peat-forming Sphagnum mosses from more than 60,000 ha of bog
30 (Lee et al., 1986). This has led to a loss of water retention and widespread erosion.
31 Nitrogen pollutants from atmospheric deposition have been implicated in this process,
August 1991
10-114 DRAFT-DO NOT QUOTE OR CITE
-------
TOO
500
300
too
number
of
species
n=1805
? X 1 3 5 7 9
poor rich
!4 of species
SO
threatened
non threat.
13579
poor rich
% fraction of
species c)
50
30
10
(x=0.20l
17=0.35)
13579
poor rich
Figure 10-20. Distribution of 2,164 Central European plant species in the gradient of
nitrogen indicator values (from Ellenberg, 1988, p. 379).
(a) "?" not known; "x" indifferent
" 1" most pronounced nitrogen deficiency
"3" poor in nitrogen
"5" just sufficient in nitrogen
"7" more often found at places rich in N
"8" nitrogen indicator -
"9" surplus nitrogen to polluted with N
"2", "4", "6" intermediate : . • .
(b) most of the threatened species can only compete on nitrogen-deficient stands (57 "potentially threatened"
species not regarded).
(c) the fraction of threatened species within the total of species in a given class of nitrogen indicator value is
deminishing with better nitrogen supply. It remains constant from value "5" upwards (see above).
1
2
3
4
5
6
7
although studies of this particular area should be interpreted cautiously because of its long
history of exposure to multiple pollutants. The combination of NO3" and NH4+ deposition,
about 32 kg N ha"1 yr"1, is more than double the deposition rates in the Berwyn Mountains in
North Wales, which still support healthy Sphagnum communities, and contributes
significantly to a supraoptimal nitrogen supply (Lee et'al., 1986). In The Netherlands there
has been a great decline during the past 3 decades in communities dominated by iosetids in
soft water areas and their conversion to later successional stages dominated by grasslands or
August 1991 10-115 DRAFT-DO NOT QUOTE OR CITE
-------
WETLAND and MOORLAND
OFTEN MECHANICALLY
DISTURBED PLACES
1 T 1
threatened n=t17~—
nan threatened n=119
1
P°cr rich poor rich
HE ATHLAND and GRASSLAND WOODLAND and BUSH
Figure 10-21. Distribution of Central European plant species along a gradient of
nitrogen indicator values (see Figure 10-20) across ecosystem types (from
Ellenberg, 1988, p. 380).
1 by Juncus bulbosus (rush) and Sphagnum spp. (Roelofs, 1983, 1986; Roelofs et al., 1984;
2 Schuurkesetal., 1986).
3 Vermeer and Berendse (1983) correlated biomass with species numbers and soil
4 chemical characteristics in several fen and grassland communities in The Netherlands. In
5 fens they found a negative correlation between biomass and NH4+ concentration and a
August 1991 10-116 DRAFT-DO NOT QUOTE OR CITE
-------
1 positive correlation between biomass and pH. There was also a positive correlation between
2 biomass and number of species. In wet grasslands a positive correlation was found between
3 biomass and NO3", Pi? and K+. In all wetland types investigated, they report that species
4 number was greatest when the standing biomass of the site was in the range of 400-500 g/m"
5 2. They concluded that domination by a few species is associated with eutrophic conditions
6 at the high end of the biomass scale as well as with conditions unfavorable for growth at the
7 low end of the scale. Similarly, in wetlands of eastern Ontario and western Quebec, the
8 greatest diversity of species (3-24 per 0.25 m2) occurs at intermediate standing crops (60-500
9 g/m"2) and the lowest density of species (2-5 per 0.25 m2) at standing crops greater than
10 1,500 g/m2 (Moore and Keddy, 1989; Wisheu and Keddy, 1989). In Great Britain species
11 density in fens was greatest (about 12 per 0.25 m2) at standing crops less than 1,000 g m"2
12 and lowest (3 per 0.25 m2) when standing crop was 4,000 g m"2 or greater (Wheeler and
13 Giller, 1982). Exceptions to this trend are found where annual mowing and harvest of
14 wetland vegetation minimize the accumulation of surface litter (Verhoeven et al., 1988), and
15 possibly where intense pressure from grazing animals favors domination by specific plant
16 species (Jensen, 1985; Berendse, 1985).
17
18 10.5.4.5 Mechanisms of Nitrogen Control Over Ecosystem Structure
19 Nitrogen supplied in excess of a plant's nutritional requirements has a direct toxic effect
20 on some species. The concentrations of six elements in the tissues of five Sphagnum species
21 have been investigated in relationship to atmospheric deposition in Europe (Ferguson et al.,
22 1984). When Sphagnum species were transplanted from a relatively clean-air site to a
23 polluted site, the concentrations of N, S, Pb, Fe, and P increased significantly, but the
24 concentration of K did not. The greatest change observed was for nitrogen which increased
25 by absolute amounts that varied from 17.7 mg per gram of tissue in Sphagnum recurvum to
26 5.3 mg/g in Sphagnum capillifolium above control levels of about 10 mg/g (1% of dry
27 weight). Since the nitrogen supply originating from the soil probably did not differ, as
28 indicated by the similarity in total nitrogen concentration of the peat from the polluted and
29 clean sites, it is possible that nitrogen deposition had a direct effect on nitrogen uptake in
30 these species. The authors concluded that the element supply from deposition at the polluted
31 site, where N deposition is 43 kg N ha"1 yr"1, is supra-optimal for growth of ombrotrophic
August 1991
10-117 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
Sphagnum species. They noted the existence of a "good" Sphagnum cover at one site where
a nitrogen deposition rate of 20 kg N ha"1 yr"1 was measured. Similarly, Press et al. (1986)
observed tissue nitrogen concentrations as high as 2.5% of dry weight in Sphagnum
cuspidatwn transplanted to a site of high N deposition in northern Britain and found that this
level of nitrogen was associated with decreased growth.
Competitive relationships among species change with the nitrogen status of the
environment. In weakly buffered ecosystems, a high deposition of NH4+ leads to
acidification and nitrogen enrichment of soil. Consequently, plant species characteristic of
poorly buffered environments disappear. Among the acid tolerant species there will be
competition between slow growing and fast growing nitrophilous grasses or grass-like species.
This process contributes to the observed change from heathlands into grasslands. Molinia
caemlea and/or Deschampsiaflexuosa (grasses) expand at the expense of Erica tetralix or
Calluna vulgaris (shrubs) and other heathland species (Berendse and Aerts, 1984; Roelofs
et al., 1987; Aerts and Berendse, 1988, 1989). In over 70 heathlands investigated, the shrub
bogs dominated by Erica tetralix or Calluna had dissolved NH4+ levels in the, soil water of
55 and 84 /iM, while those dominated by the grasses Deschampsia and Molinia had average
NH4+ concentrations of 248 and 429 /tM (Roelofs et al., 1987).
Several controlled growth studies also have been carried out to identify the mechanisms
of nitrogen control over species composition. This is a non-trivial task since there are a great
number of interactions among biochemical and geochemical processes. There are direct and
indirect effects of nitrogen deposition, and cause and effect can be difficult to ascertain.
Roelofs (1986), for example, states that acidification, which can result from deposition of
either NOX, SO42", or NH4+, can decrease the availability of dissolved CO2 in water which
leads to the complete elimination of submerged plant species. Deposition of NH4+ and its
subsequent nitrification or absorption by plants generates acidity. Biochemical conversions of
SO4 " and NO3" generate alkalinity. These processes are mediated by bacteria, macrophytes,
and algae (Kelly et al., 1982; Raven, 1985). Atmospheric deposition of nitrogen can
significantly affect the nitrogen budget of some wetland ecosystems, their acidity, and their
carbon budgets (Roelofs, 1986).
Schuurkes et al. (1986) studied effects of acidification and nitrogen supply on growth of
several common wetland plants under controlled laboratory conditions. All species utilized
August 1991
10-118 DRAFT-DO NOT QUOTE OR CITE
-------
1 NH4+ and NO3~ as a nitrogen source except Sphagnum flexuosum, which did not assimilate
2 NO3". When NH4rf and NO3" were offered simultaneously in equal amounts, NO3" uptake
3 was the dominant form of nutrition (63-73%) in plants that are characteristic of soft waters
4 (low Ca2+ and Mg2+), while NH4+ strongly dominated the nutrition (85-90%) in species
5 from acid waters. Differences in the site of uptake, either leaves or roots, among species
i O
6 were also found. They concluded that high deposition of NH4 and SO4 ", the most
7 important sources of acidification in The Netherlands, is leading to an expansion of acid-
8 tolerant nitrophilous plants.
9 The nutrition of Sphagnum is apparently species specific. Although S. flexuosum did
10 not assimilate NO3" (Schuurkes et al., 1986), the activity of nitrate reductase in S.
11 cuspidatum (Press and Lee, 1982) and mS.juscum (Woodin et al., 1985) clearly shows that
12 NO3~ can be utilized by these species. S. magellanicum was shown to grow best when given
13 the equivalent of 4.1 kg NO3"-N ha"1 yr"1 plus 19 kg NH4+-N ha"1 yr"1 in simulated rain;
14 when given 0.25X that amount of NO3" and 1.5X (and 4X) as much NH4+ growth decreased
15 (Rudolph and Voigt, 1986). Bayley et al. (1987) reported that the dominant Sphagnum spp.
16 in a poor fen in Ontario, S. juscum and S. magellanicum, were able to assimilate a NO3"
17 input of 4.71 kg N ha"1 yr"1, including 1.6 kg N ha"1 yr"1 applied in simulated acid rain, and
18 growth increased at least during the first year after the additional nitrogen was applied.
19 Roelofs et al. (1984) observed that growth of S. cuspidatum was greatest in a medium
20 containing 500 0M NH4+, and less at 1,000 or 100 /tM NH4+, while Press et al. (1986)
21 observed that the best growth of this same species occurred in N-free solutions, and that even
22 small additions (10 fJ-M) of either NH4+ or NO3" reduced growth. It is doubtless that some
23 variations in results of nutritional studies are influenced by other variables, like pH.
24 The genus Sphagnum is an important group in bogs everywhere, and it is important to
25 understand its nutritional physiology and ecology. However, it should be emphasized that the
26 consequences of nitrogen fertilization in a natural environment, with fluctuating climate and
27 competition among numerous species, can be quite different from what may be predicted
28 from studies of a single species in laboratory culture. For example, Aerts et al. (1989) assert
29 that competition for light dictates the outcome of competition between species that differ in
30 growth rate potential and nutrient requirement.
August 1991
10-119 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
In a two-year greenhouse experiment designed to differentiate between acid and nitrogen
effects, Schuurkes et al. (1987) exposed mixtures of different wetland plant species to
simulated rain containing various combinations of SO42", NH4+, and NO3". Marked changes
were observed in systems receiving rain with 510 and 1,585 #M NH4+; plants typical of
nutrient-poor soft waters, like the isoetids Littoretta uniflora (shoreweed), Luronium natans
(water plantain) and Pilularia globulifem, were adversely affected at this level of nitrogen
input, while other species (Juncus bulbosus, Sphagnum cuspidatum, and the grass Agrostis
canind) expanded. Acidification with none or only a small NH4+ addition had no clear
effects, although biomass of Sphagnum was slightly higher. Within H2SO4 treatments, only
pH 3.5 rain markedly acidified the water. Based on these experiments, Schuurkes et al.
(1987) recommended that to preserve the remaining oligotrophic wetlands, acid inputs should
not exceed 250 mol ha"1 yr"1, and that the annual nitrogen deposition should not be greater
than 1,380 mol ha"1 yr"1 or 19.4 kg N ha"1 yr"1 (NO3" + NH4+), except that the potential
acidifying influence of this nitrogen input, if in the form of NH4+, exceeds the allowable
acid input. This limit is supported by Liljelund and Torstensson (1988) -who concluded from
their review that the limit for many species may be well below 20 kg N ha"1 yr"1 and for
oligotrophic (nutrient-poor) bogs is probably about 10 kg N ha"1 yr"1. These limits are
exceeded currently in the United States where wet nitrate deposition alone exceeds 15 kg
N ha"1 yr"1 over most of the midwest, New York, and New England (Zemba et al., 1988).
10.6 AQUATIC EFFECTS OF NITROGEN OXIDES
.10.6.1 Introduction
For a variety of reasons, nitrogen deposition has not historically been considered a
serious threat to the integrity of aquatic systems. Most terrestrial systems have been assumed
to retain nitrogen strongly, leading to a small probability that deposited nitrogen will ever
make its way to the surface waters that drain these terrestrial systems. Nitrogen within
aquatic ecosystems can arise from a variety of sources, including point-source and
non-point-source pollution and biological fixation of gaseous nitrogen, in addition to the
deposition of nitrogen oxides. In cases where nitrogen is known to be affecting aquatic
systems, it has been assumed that some source other than deposition is responsible. The
August 1991
10-120
DRAFT-DO NOT QUOTE OR CITE
-------
1 amounts of nitrogen provided to aquatic systems by these other sources often outweigh by a
2 large margin the amount of nitrogen potentially provided by atmospheric deposition. In the
3, past decade, however, our understanding of the transformations that nitrogen undergoes
4 .within watersheds has increased greatly, and in areas of the country where non-atmospheric
5 , sources of nitrogen are small, we can begin to infer cases where nitrogen deposition is having
6 an impact on aquatic systems.
7 Estimating the effects of nitrogen oxide emissions and nitrogen deposition on aquatic
8 , systems is made difficult by the large variety of nitrogen compounds found in air, deposition,
9 watersheds and surface waters, as well as the myriad of pathways through which nitrogen can
10 • be cycled in terrestrial and aquatic ecosystems. These complexities have the effect of
11 de-coupling nitrogen deposition from nitrogen effects, and reduce our ability to attribute
12 known aquatic effects to known rates of nitrogen deposition. The organization of this section
13 reflects this complexity. Because an understanding of the ways that nitrogen is cycled
14 through watersheds is critical to our understanding of nitrogen effects, the section begins with
15 a brief description of the nitrogen cycle, and of the transformations 'of nitrogen that may
16 occur in watersheds. Each of the known possible effects of nitrogen deposition (acidification,
17 eutrophication and direct toxicity) is discussed separately. Within these discussions, evidence
18 for the importance of nitrogen in causing observed effects is discussed separately from
19 evidence that deposition is the source of the nitrogen observed in affected systems.
20
21 10.6.2 The Nitrogen Cycle
22 Atmospheric nitrogen can enter aquatic systems either as direct deposition to water
23 surfaces, or as nitrogen deposition to the terrestrial portions of a watershed (Figure 10-22).
24 Nitrogen deposited to the watershed is then routed (e.g., through plant biomass and soil
25 microorganisms) and transformed (e.g., into other inorganic or organic nitrogen species) by
26 watershed processes, and may eventually run off into aquatic systems in forms that are only
27 indirectly related to the original deposition. Much of the challenge of determining when
28 nitrogen deposition is having an effect on aquatic systems depends on our ability to track
29 nitrogen -on its path through watersheds. In most cases, this tracking cannot be accomplished
30 outside of a carefully controlled research program, and we are forced to make educated
31 guesses about the likelihood that the nitrogen observed in aquatic systems was originally of
August 1991
10-121 DRAFT-DO NOT QUOTE OR CITE
-------
DEPOSITION
WET
NOJ NH4*
DRY
NOX'NHX
..h.
assimilation N.
nitrification
{ assimilation
mineralization,
DEAD
ORGANIC
MATTER
7
M1CR08IAL
BIOMASS
nitrogen
fixation
denitrification
LEACHING
WATER
Figure 10-22. A simplified watershed nitrogen cycle. Only the major pathways are
shown. The boxes represent major pools of nitrogen in terrestrial
ecosystems, and the lines represent the major pathways and processes
affecting nitrogen transformations. The wavy line represents the soil
surface. Modified from Skeffington and Wilson (1988).
1
2
3
4
5
6
atmospheric origin. The strength of these educated guesses will depend, to a large degree, on
our ability to identity which nitrogen transformations are occurring and which are not. By
eliminating other possible sources or sinks of nitrogen, we are in a stronger position to
determine in which cases observed nitrogen effects are caused indirectly by atmospheric
deposition. Our understanding of the nitrogen cycle in terrestrial and aquatic ecosystems
therefore plays a central role in controlling our understanding of deposition effects. The key
August 1991
10-122 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
elements of the nitrogen cycle, particularly those that are thought to be important in
determining whether atmospherically-derived nitrogen will have an effect on aquatic systems,
are discussed briefly in this section.
10.6.2.1 Nitrogen Inputs
Watersheds are generally several orders of magnitude larger than the surface waters that
drain them, and so the majority of the atmospheric deposition that may potentially enter
aquatic systems falls first on some portion of the watershed. Nitrogen may be deposited to
the watershed, or directly to water surfaces, in a variety of forms, including nitrate (NO3"),
ammonium (NH4+), and organic nitrogen in wet and dry deposition. In addition, plants may
absorb gaseous nitrogen (as NOX; Rowland et al., 1985) or nitric acid vapor (Vose et al.,
1989), and nitrogen thus absorbed may subsequently enter the watershed nitrogen budget as
litter fall, or through the death of plant biomass (Parker, 1983; Olson et al., 1985). These
nitrogen constituents are the same as those comprising direct deposition to terrestrial
ecosystems recently described by Lindberg et al. (1986) also Section 10.4.
Concentrations of NO3" and NH4+ in precipitation vary widely throughout North
America, depending largely on the proximity of sampling sites to sources of emissions.
Galloway et al. (1982) report mean concentrations of NO3~ and NH4+ of 2.4 /zeq • L"1 and
2.8 #eq • L"1, respectively, for a site in central Alaska. In the Sierra Nevada Mountains of
California, mean concentrations of NO3" and NH4+ for the period 1985-1987 were 5.0 and
5.4 /ieq • L"1, respectively (Williams and Melack, in press a). In a comparison of nitrogen
deposition at lake and watershed monitoring sites in the northern United States and southern
Canada, Linsey et al. (1987), found NO3" concentrations ranging from 15 to 40 #eq • L"1 and
NH4+ concentrations from 10 to 50 /*eq . L"1 in areas considered remote but influenced by
prairie dust and long range acidic deposition; neither ion dominated over the other. In some
areas closer to anthropogenic nitrogen sources (e.g., in northeastern United States and
southeastern Canada) volume-weighted mean NO3" concentrations range from 30 #eq • L"1
(e.g., in the Adirondack and CatsMll mountains of New York) to 50 jLteq • L"1 (e.g., in the
eastern Great Lakes region), while mean NH4+ concentrations range from 10 to 20 #eq • L"1
in the same areas (Stensland et al., 1986). Ammonium concentrations are highest
(ca. 40 #eq • L"1) in the agricultural areas of the mid-western United States.
August 1991
10-123 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
Deposition of nitrogen will depend on the concentration in precipitation, the volume of
water falling as precipitation, and the amount of nitrogen in dry deposition (see
Section 10.4 of this report; see also Sisterson et al., in press). The last of these values (dry
deposition) is difficult to measure, and is often estimated as a fraction (e.g., 30-40%) of wet
deposition (Baker, 1991). Given the range of concentrations mentioned in the previous
paragraph, and the volumes of precipitation falling in different regions of North America,
estimates of nitrogen deposition rates range from less than 12 eq • ha"1 • yr"1 in Alaska to
near 800 eq • ha"1 • yr'1 in the northeastern United States (Table 10-20). !
f".
TABLE 10-20. RATES OF NITROGEN DEPOSITION IN SEVERAL
AREAS OF NORTH AMERICA
Area
Alaskaa
(Poker Flat)
Sierra Nevada, CAb
(Emerald Lake)
Ontario, Canada0
(Experimental Lakes Area)
British Columbia, Canada0
Upper Midwestd
Southeastern U.S.e
(Walker Branch, TN)
New Hampshire0
Catskillsa
Adirondacksd
eq
NO3"
6.9
79
125
260
300
540
464
580
590
• ha"1 • yr"1
NH4+
4.8
85
140
130
210
180
200
292
190
Total
11.7
164
265
390
510
720
664
874
780
Source
Galloway et al. (1982)
Williams and Melack (in press, a)
N JT 7 /
Linsey et al. (1987)
Feller (1987)
Driscoll et al. (1989a)
Lindberg et al. (1986)
Likens (1985)
Stoddard and Murdoch (1991)
V S
Driscoll et al. (1989a)
"Dry deposition estimated as 35% of total deposition.
bDry deposition sampled as part of sno'yvpack; no correction for dry deposition made.
"Bulk precipitation measurements; no correction for dry deposition made.
dValues corrected for dry deposition based on ratios in Hicks (1989).
"Includes estimates for dry deposition and gaseous uptake of nitrogen areas, DON can occur in greater
concentrations than the inorganic species (Moore and Nuckols, 1984).
1
2
Generally NO3" dominates over NH4+ at sites close to emission sources (Linsey et al.,
1987; Altwicker et al., 1986). Dissolved organic nitrogen (DON) concentrations are highly
August 1991
10-124 DRAFT-DO NOT QUOTE OR CITE
-------
1 variable in precipitation but often amount to 25% to 50% of inorganic nitrogen deposition
2 values (Linsey et al., 1987; Manny and Owens, 1983; Feller, 1987).
3 .. .
4 10.6.2.2 Transformations
5 Because the majority of nitrogen deposition falls first on some portion of the watershed,
6 the transformations that nitrogen undergoes within the watershed (e.g., in soils, by microbial
7 . action, and in plants) will play a major role in determining what forms and amounts of
8 nitrogen eventually, reach surface waters. Much of the following discussion is therefore
9 focused on terrestrial processes that alter the forms and rates of nitrogen supply. It is these
10 processes that, to a large degree, determine whether nitrogen deposition will ever reach lakes,
11 streams and estuaries, and therefore they are very important in controlling the effects of
12 nitrogen deposition. Many of these same processes occur also within surface waters, and a
13 specific discussion of these processes, and their importance, follows the discussion of
14 nitrogen transformations.
15 ,
16 Nitrogen Assimilation
17 Nitrogen assimilation is the uptake and metabolic use of nitrogen by plants
18 (Figure 10-22). Assimilation by both terrestrial and aquatic plants will play a role in
19 determining whether nitrogen deposition affects aquatic systems. Assimilation by the
20 terrestrial ecosystem controls the form of nitrogen eventually released into surface waters, as
21 well as affecting the acid/base status of soil and surface waters. Terrestrial assimilation is a
22 major form of nitrogen removal in watersheds, and may in fact be sufficient to prevent all
23 atmospherically-derived nitrogen from reaching surface waters (Vitousek and Reiners, 1975).
24 Nitrogen is the most commonly limiting nutrient in forest ecosystems in North America
25 (Cole and Rapp, 1981). Because the primary use of nitrogen in plant biomass is the
26 formation of amino acids, and reduced nitrogen is the most energetically favorable form of
27 nitrogen for incorporation into amino acids, uptake of NH4+ is generally favored over uptake
28 of NO3" by terrestrial plant species. This demand for NH4+ over NO3" contributes to the
29 common pattern of low NH4+ concentrations in waters draining forested watersheds in the
30 United States. The form of nitrogen used by terrestrial ecosystems effects strongly the
31 acidifying potential of nitrogen deposition (Figure 10-23). Ammonium uptake is an
August 1991
10-125 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
acidifying process (i.e., uptake of NH4+ releases one mole of hydrogen per mole of nitrogen
assimilated):
NH4+ + R • OH = R • NH2 + H2O + H+ (Eq. 10.6-1)
The biological uptake of NO3~, on the other hand, is an alkalinizing process (i.e., uptake of
NO3" consumes one mole of hydrogen per mole of nitrogen assimilated):
R • OH + N03- + H+ = R • NH2 + 2O2
(Eq. 10.6-2)
Nitrification
Nitrification is the oxidation of ammonium (NH4+) to nitrate (NO3\), and is mediated
by bacteria and fungi in both the terrestrial and aquatic portions of watersheds. It is an
important process in controlling the form of nitrogen released to surface waters by
watersheds, as well as in controlling the acid/base status of surface waters (Figure 10-22).
Nitrification is a strongly acidifying process, producing two moles of hydrogen for each mole
of nitrogen (NH4+) nitrified (Figure 10-23):
NH+
2Q =
2H
+
H2O
(Eq. 10.6-3).
Because nitrification in forest soils commonly transforms NH4+ into NO3", the
acidifying potential of deposition (attributable to nitrogen) is often defined as the sum of
NH4+ and NO3", assuming that all nitrogen will leave the watershed as NO3" (e.g., Hauhs
etal., 1989).
In most soUs, nitrification is limited by the supply of NH4+ (Likens et al., 1970;
Vitousek et al., 1979), creating a high demand for NH4+ on the part of nitrifying soil
microbes. This microbial demand for NH4+, coupled with the demand for NH4+ on the part
of terrestrial plants (discussed above) leads to surface water concentrations of NH4+ which
are almost always unmeasurable. Nitrification rates may also be limited by inadequate
microbial populations, lack of water, allelopathic effects (toxic effects produced by inhibitors
manufactured by vegetation), or by low soil pH. Of these other potential limiting factors,
soil pH plays an obviously vital role in any discussion of the acidification of surface waters
August 1991
10-126 DRAFT-DO NOT QUOTE OR CITE
-------
IN IN
wor NH4*
-1
IN
IN
ORGANIC
MATTER
J3 -1
-1
DEPOSITION, FERTILISERS
PLANTS
PENITRIFIERS
./DECOMPOSITION •
FIGURES REPRESENT H
TRANSFERS TO THE
SOIL OR WATER
OUT OUT
Figure 10-23. The effect of nitrogen transformations on the watershed hydrogen ion
budget. One hydrogen ion is transferred to the soil solution or surface
water (+1) or from the soil solution or surface water (-1) for every
molecule of NO3" or NH4+ which crosses a compartment boundary. For
example, nitrification follows the pathway for NH4+ uptake into organic
matter (+1), and is leached out as NO3" (+1), for a total hydrogen ion
production of +2 for every molecule of NO3" produced. Modified from
Skeffington and Wilson (1988).
1
2
3
4
5
6
7
8
9
10
11
12
by nitrogen deposition. Nitrification has traditionally been thought of as an acid-sensitive
process (Driscoll and Schaefer, 1989; Aber et al., 1989), but high rates of nitrification have
been reported from very acid soils (i.e., pH <4.0) in the northeastern United States
(Vitousek et al., 1979; Novick et al., 1984; Rascher et al., 1987) and in Europe (van
Breemen et al., 1982). In the southeastern United States, Montagnini et al. (1989) were
unable to find any effect of pH on nitrification, or to stimulate nitrification by buffering acid
soils. In a survey of sites across the northeastern United States, McNulty et al. (1990) found
no correlation between nitrification rates and soilpH, but a strong association (r2 = 0.77)
with rates of nitrogen deposition. The weight of evidence suggests that nitrification will
proceed at low soil pH values as long as the supply of NH^4" is sufficient.
August 1991
10-127 DRAFT-DO NOT QUOTE OR CITE
-------
1 Denitrification
2 Denitrification is the biological reduction of NO3" to produce gaseous'forms of reduced
3 nitrogen (N2, NO, or N2O) (Payne, 1981). Denitrification is an anaerobic process (i.e., it
4 proceeds only in environments where oxygen is absent) whose end product is lost to the
5 atmosphere (Figure 10-22). In terrestrial ecosystems, denitrification occurs in anaerobic
6 soils, especially boggy, poorly drained soils, and has traditionally been considered a relatively
7 unimportant process outside of wetlands (Post et al., 1985). It has been suggested, however,
8 that denitrification could be an episodic process, occurring after such events as spring snow
9 melt and heavy rain storms, when soil oxygen tension is reduced (Melillo et al., 1983). No
10 single equation can describe the denitrification reaction, because several end-products are
11 possible. However, denitrification is always an alkalinizing process, consuming one mole of
12 hydrogen for every mole of nitrogen denitrified (Figure 10-23). Denitrification can be
13 involved in the production or consumption of N2O, a product which may have considerable
14 significance as a greenhouse gas (Matson and Vitousek, 1990; Halm and Crutzen, 1982). In
15 a review of the effects of acidic deposition on denitrification in forest soils, Klemedtsson and
16 Svensson (1988) conclude that denitrification rates are often limited by the availability of
17 oxygen, and may therefore be relatively insensitive to increases in nitrogen deposition. It has
18 been suggested that the production of N2O may increase in acidified soils (Knowles, 1982),
19 but few field data are available to test this idea. Rates of N2O production in soil waters have
20 been shown to increase markedly after forest clear-cutting (Bowden and Bormann, 1986;
21 Melillo et al., 1983), and in areas of both high nitrogen deposition, and intensive forest
22 management, N2O production may be of concern. Nitrous oxide production is strongly
23 influenced by soil temperature, soil NO3" concentration, and soil moisture; Davidson and
24 Swank (1990) suggest that one or more of these factors may commonly limit N2O production
25 in natural systems.
26
27 Nitrogen Fixation
28 Gaseous atmospheric nitrogen (N^ can be fixed to produce NH4+ by a wide range of
29 single-celled organisms including blue-green algae (cyanobacteria), and various aerobic and
30 anaerobic bacteria. Symbiotic nitrogen-fixing nodules are present on the roots of some early
31 successional forest species (Boring et al., 1988). In headwater streams, nodules on rooting
August 1991
10-128
DRAFT-DO NOT QUOTE OR CITE
-------
1
•2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
structures of riparian vegetation (e.g., Alnus sp.) can also be important nitrogen fixers
(Binkley, 1986). Ordinarily, nitrogen fixation has no direct effect on the acid/base status of
soil or surface waters:
N2 + H2O •¥ 2R • OH = 2R • NH2 +' 3/iO2
(Eq. 10.6-4)
Nitrogen fixation in excess of biological demand, however, can lead to nitrification or
mineralization of organic nitrogen, and ultimately lead to acidification of soil or surface
waters (Franklin et al., 1968; Van Miegroet and Cole, 1985).
Mineralization
Mineralization is the bacterial decomposition of organic matter, releasing NH4+ that
can subsequently be nitrified to NO3~. Mineralization is an important process in watersheds,
as it recycles nitrogen that would otherwise be lost from the system through death of plants,
'or as leaf litter (Figure 10-22). In a comparative study of mineralization in soils,
Nadelhoffer et al. (1985) found nitrogen mineralization rates ranging from 3,600 to
7,600 eq • ha^-yr'1 under deciduous tree species, and from 2,300 to 4,700 eq • ha"1 • yr"1
under coniferous species. These rates should be compared to nitrogen deposition rates of
400-800 eq • ha"1 • yr"1 for high deposition areas of the Northeast. Nadelhoffer et al. (1985)
also report estimated rates of nitrogen uptake that were 20 'to 80% higher than rates of
mineralization, suggesting that mineralization can supply the majority, but not all, of the
nitrogen needed for plant growth in these forests.
The effect of mineralization on the acid/base status of draining waters will depend on
the form of nitrogen produced. The conversion of organic nitrogen (e.g., from leaf litter) to
NH4+ consumes one mole of hydrogen per mole of nitrogen produced (Figure 10-23), and
can be thought of as the reverse of the reaction in Equation 10.6-1. Organic nitrogen which
is mineralized and subsequently oxidized (nitrified) to NO3" (Equation 10.6-3) produces a net
of one mole of hydrogen per mole of NO3" produced. Because the production of organic
nitrogen (i.e., assimilation) can either produce or consume hydrogen (depending on whether
NO3" or NH4+ is assimilated), the net (ecosystem) effect of mineralization depends both on
the species entering the watershed, and on the species'leaving the watershed (Figure 10-23).
August 1991
10-129
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
In ecosystems where plant growth is limited by the availability of nitrogen, „
mineralization is also limited by nitrogen, in the sense that additions of nitrogen to the leaf
litter will speed decay, and increase the rate at which nitrogen is immobilized by
decomposers (Melfflo et al., 1984; Taylor et al., 1989). Nitrogen limitation of
decomposition is in part due to the low nitrogen content typical of litter, resulting from the
retranslocation of nitrogen out of leaves during senescence.
10.6.2.3 Nitrogen Saturation
Much of the debate over whether aquatic systems are being affected by nitrogen
deposition centers on the concept of nitrogen saturation of forested watersheds. .Nitrogen
saturation can be defined as a situation where the supply pf nitrogenous compounds, from the
atmosphere exceeds the demand for these compounds on the part of watershed plants and
microbes (Skeffington and Wilson, 1988; Aber et al., 1989). Under conditions of nitrogen
saturation, forested watersheds that previously retained nearly all of nitrogen inputs, due to a
high demand for nitrogen by plants and microbes, begin to supply more nitrogen to the
surface waters that drain them.
The major aspects pf the nitrogen saturation concept can be described by a simple
analogy. In this analogy, forested watersheds are likened to, a dry sponge. When a sponge is
dry, it has a high capacity to absorb water deposited on its surface, just as a
nitrogen-deficient forest has a high capacity to accumulate nitrogen from deposition. The
capacities of both the forest and the sponge are related to the rate of supply of nitrogen or
water, respectively, at least in the early stages of nitrogen or water addition. .That is, if we
pour water slowly onto the sponge it will retain nearly all of it, but if the water is poured
quickly the sponge will retain most of the water, while some leaks out. This same
characteristic in forests results in the common seasonal pattern for streams that drain forested
watersheds. During seasons of low nitrogen deposition (e.g., during dry seasons) nearly all
of the deposited nitrogen is retained in soils and forest biomass, and stream water
concentrations of nitrogen are low or undetectable. When nitrogen deposition is at a high
rate, however, it may exceed (temporarily) the capacity of the system to absorb nitrogen, and ,
some will leak into the streams that drain the watersheds. This simplified pattern of low
August 1991
10-130 DRAFT-DO NOT QUOTE OR CITE
-------
1 nitrogen concentrations in streams during dry seasons, and high concentrations during wet
2 seasons, is a very common one.
3 Returning to the sponge analogy, we know that if we add water to the sponge at a
4 constant rate Over some period of time, eventually the sponge will become saturated; its
5 capacity to absorb water will be very low, and the rate of water leakage from the sponge will
6 depend on the rate of water addition. In a similar manner, the capacity of forested
7 ecosystems to retain nitrogen may be exceeded by long-term high rates of nitrogen
8 deposition. At some critical point in this process, forested watersheds will begin chronically
9 "leaking" nitrogen to the streams that drain them. This'is the phenomenon commonly
10 thought of as nitrogen saturation. The major difference between this simple sponge analogy
11 and the cycling of nitrogen in nitrogen-deficient forests is that the forests have
12 nitrogen-retaining capacities (i.e., biological demands for nitrogen) which vary with season,
13 depending on whether forest growth is occurring or not (i.e., nitrogen retention is higher
14 during the growing season). One major effect of this complexity is to amplify the seasonal
15 differences in nitrogen leakage. This amplification results when the growing season
16 coincides, as it does in most forested areas, with a period of lower deposition (late spring and
17 summer); many forests are dormant during the seasons of highest atmospheric deposition
18 (winter and early spring) and so both supply and demand favor the leakage of some nitrogen
19 from the system.
20 Aber et al. (1989) have proposed a hypothetical time course for a watershed response to
21 'chronic nitrogen additions (Figure 10-24), describing both the changes in nitrogen cycling
22 that are proposed to occur, as well as the plant responses to changing levels of nitrogen
23 availability. The four stages (Figure 10-24), correspond to those described by Smith (1974)
24 and Bormann (1982) for ecosystem response to pollution loading. Stage 0 is the
25 pre-treatment condition. In Stage 1, increased deposition is occurring, but effects on the
26 ecosystem are not evident. Many forested watersheds in the United States would be
27 considered to exist at this stage. For a limiting nutrient such as nitrogen, a fertilization effect
28 might result in increased ecosystem production and tree vigor at Stage 1. Retention of
29 nitrogen is very efficient, and little or no nitrogen would be lost to surface waters that drain
30 Stage 1 watersheds.
August 1991
10-131
DRAFT-DO NOT QUOTE OR CITE
-------
c
•I
N Mineralization
Nitrification
N Inputs
NO 3
Leaching
N2O
Emission
Additions
Begin
Stage 0 1
l I
Saturation Decline
NPP
Foliar Biomass
Foliar N Concentration
Fine Root Mass
Nitrate Assimilation
Stage
i
Additions
Begin
0
l l
Saturation Decline
1
Figure 10-24. Hypothetical time course of forest ecosystem response to chronic nitrogen
additions. Top: relative changes in rates of nitrogen cycling and nitrogen
loss. Bottom: relative changes in plant condition (e.g., foliar biomass and
nitrogen content, fine root biomass) and function (e.g, net primary
productivity (NPP) and nitrate assimilation) in response to changing levels
of nitrogen availability. From Aber et al. (1989).
August 1991
10-132
DRAFT-DO NOT QUOTE OR CITE
-------
1 In Stage 2 of the Aber et al. (1989) hypothetical time course, negative effects occur, but
2 they are subtle, nonvisual, and/or require long time scales to detect. Stage 2 could be
3 considered to correspond to the "damp" stage of the sponge analogy, where rapid rates of
4 water addition will cause leakage; if nitrogen is added over short time scales (e.g., in storms)
5 to Stage 2 watersheds, leakage of NO3" would result. Only in Stage 3 do visible effects on
6 the forests occur, resulting in major environmental impacts. Aber et al. (1989) emphasize
7 that different species and environmental conditions could alter the timing of effects illustrated
8 in Figure 10-24).
9 The high capacity of forested watersheds to retain nitrogen is responsible for the concept
10 that nitrogen deposition does not affect aquatic ecosystems. However, as nitrogen saturation
11 proceeds, the ability of watersheds to retain nitrogen decreases, and the potential for nitrogen
12 effects in aquatic systems grow. One of the major tasks in assessing nitrogen effects on
13 surface waters is the recognition of the early stages of nitrogen saturation. It has been
14 suggested (Grennfelt and Hultberg, 1986; Driscoll and Schaefer, 1989; Stoddard and
15 Murdoch, 1991) that chronic nitrogen leakage from forested watersheds (Stage 3 in
16 Figure 10-24) will be preceded by an amplification of the seasonal pattern in surface water
17 nitrogen. The rationale behind this suggestion stems from the idea that biological activity is
18 probably not limited be a single nutrient or physical factor year-round. Forest growth, for
19 example, is probably limited by temperature during the winter months (Likens et al., 1977;
20 Vitousek, 1977), and passes through a phase in the spring when growth is limited either by
21 temperature or by nitrogen, depending on whether nitrogen deficiency or cold is more severe
22 at any given point in time. As forests become more nitrogen sufficient (i.e., as nitrogen
23 saturation proceeds), the period of temperature limitation is likely to be extended further and
24 further into the spring season, or to be replaced by a period of limitation by some other
25 nutrient or physical factor. This longer period of low nitrogen demand (on the part of
26 forests) is likely to produce larger or longer periods of elevated nitrogen concentrations in
27 lakes and streams during spring snow melt. This extended period of low demand for nitrogen
28 need not extend very far into the spring snowmelt season in order to cause nitrogen episodes,
29 as soil water charged with nitrogen may be quickly transported to soils zones below the
30 rooting zone (i.e., below the soil level where biological uptake can affect the nitrogen
31 concentration) by the movement of water during snow melt (Murdoch and Stoddard, in
August 1991
10-133
DRAFT-DO NOT QUOTE OR CITE
-------
1 review). A number of other mechanisms can produce similar results, but there are few data
2 to support their existence. These mechanisms include:
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
• An increase in the length or severity of the winter. By delaying the onset of snow
melt, longer winters would allow a larger accumulation of nitrogen in the snow pack,
and could prolong the period of temperature limitation in forests, lowering the
demand for nitrogen as the dormant period is ending. Limited available evidence,
however, suggests that winters have become shorter rather than longer, and that snow
melt is occurring earlier in the year now than it did 20 years ago (Schindler et al.,
1990).
• Prolonged defoliation of forest trees (e.g., by insect outbreaks) could lower the forest
demand for nitrogen and potentially increase NO3" concentrations in runoff water
during the growing season. Defoliation should have little effect on snowmelt NO3"
concentrations, however, because these episodes occur largely in deciduous forest
that are devoid of leaves during the dormant (winter) season.
• Increases in nitrogen deposition. These could result in more severe N03" episodes,
independent of any change in nitrogen saturation, by producing a larger accumulation
of nitrogen in winter snow packs. Because most snow melts while the forest is still
largely dormant, an increase in nitrogen storage in the snowpack could lead to an
increase in the severity of snowmelt NO3" episodes, without any long-term change in.
nitrogen retention occurring. There is no evidence, however, that nitrogen
deposition is increasing anywhere in North America (Simpson and Olsen, 1990;
Bowersox et al., 1990). Estimates of nitrogen deposition calculated from. NOX
emissions in the Northeast suggest that nitrogen deposition increased more-or-less
linearly from the turn of the century until 1970, but has remained relatively constant
in the past 20 years (Husar, 1986). The longest available records of nitrogen
deposition for a site in the Northeast (Hubbard Brook Experimental Forest) suggest
that rates of nitrogen deposition peaked in the early 1970s and have either decreased
(NH4+) or remained constant (NO3~) since then (Likens et al., 1984; Hedin et al.,
1987).
A change in the ability of forests to retain nitrogen remains, therefore, the most likely
mechanism producing increases in snowmelt NO3" episodes. According to this scenario, the,
earliest symptoms of nitrogen saturation of soils may be most visible as increases in the
severity of nitrogen episodes in the surface waters.
August 1991
10-134
DRAFT-DO NOT QUOTE OR CITE
-------
1 10.6.2.4 Processes Within Lakes and Streams
2 All of the transformations and processes discussed above (primarily in the context of
3 terrestrial ecosystems) also take place in lakes, streams and estuaries. The emphasis on the
4 transformations that occur in the watershed, before nitrogen reaches surface waters, results
5 from the necessity to establish a linkage between nitrogen deposition and nitrogen effects in
6 aquatic systems, but should not be taken to suggest that nitrogen transformations within
7 aquatic systems are of minor importance in the nitrogen cycle. In a very real sense, nitrogen
8 cycling within the terrestrial ecosystems controls whether nitrogen deposition will reach
9 aquatic systems (and in what concentrations), while nitrogen cycling within lakes, streams
10 , and estuaries controls whether the nitrogen will have any measurable effect.
11 Assimilation by aquatic plants is a key process in the potential eutrophication of surface
12 waters by nitrogen, and may also play a role in their acid/base status. The following
13 discussion of nitrogen assimilation in aquatic systems will deal mainly with the algal and
14 microbial community in phytoplankton (microscopic algal and bacterial species suspended in
15 the water column) and periphyton (algal species growing attached to surfaces). Although
16 macrophytes (macroscopic algal species) are also important in the assimilation of nitrogen,
17 the biomass of phytoplankton and smaller microbes is potentially most reactive to changes in
18 nitrogen supply. Algal uptake is a major component of the eutrophication process, and forms
19 the basis of trophic production in streams and lakes. It can also play a large role in the
20 , acid/base status of lakes. Uptake of NC)3" in lakes is an alkalinizing process, consuming one
21 mole of hydrogen per mole of nitrogen assimilated (Kelly etal., 1990).
22 . Like terrestrial plants, aquatic plants favor the uptake of NH4+ over the uptake of
23 NO3"; NH4+ uptake is energetically favorable because NO3" must first be reduced before it is
24 physiologically available to algae (Reynolds, 1984). In some circumstances organic forms of
25 nitrogen are also available for uptake by aquatic plants (reviewed by Healey, 1973). The
26 preferences by algae for the different forms of nitrogen can be related to the history of
27 availability of nitrogen species. In some algal species, the synthesis of the enzyme (nitrate
28 reductase) required to utilize a NO3" pool can be induced by high concentrations of NO3" in
29 the absence of NH4+ (Healey, 1973). The production of nitrate reductase appears to be
30 repressed by the presence of NH4+ (Eppley et al., 1979).
August 1991
10-135
DRAFT-DO NOT QUOTE OR CITE
-------
1 The potential uptake rate of inorganic nitrogen is related to ambient inorganic nitrogen
2 concentrations (e.g., Syrett, 1953), that is, cells transferred from nitrogen-deficient media to
3 nitrogen-sufficient media show higher rates of uptake than cells that are grown and remain in
4 nitrogen-sufficient media. McCarthy (1981) summarized several studies which consistently
5 showed that potential (saturated) NH4+ uptake rates were greatly enhanced in
6 nitrogen-deficient cells. This relationship is now used along with various other indices as a
7 basis to identify the degree of nitrogen limitation in phytoplankton (Vincent, 1981; Suttle and
8 Harrison, 1988). Under nitrogen-replete conditions, saturated uptake rates are low but
9 increase with increasing nitrogen deficiency.
10 A crucial difference between aquatic and terrestrial ecosystems with respect to nitrogen
11 is that nitrogen additions do not commonly stimulate growth in aquatic systems, as seems to
12 be the case in terrestrial systems, and nitrogen limitation may in fact be the exception in
13 aquatic systems rather than the rule. Determining whether nitrogen limitation is a common
14 occurrence in surface waters will play a large role in determining whether nitrogen deposition
15 affects the trophic state of aquatic ecosystems.
16 The effects of nitrogen supply on uptake and growth rates in phytoplankton and
17 periphyton is the subject of volumes of literature, a summary of which is beyond the scope of
18 this section. However, certain aspects of the limitation of algal growth by the supply of
19 nitrogen and other nutrients will be discussed later as it relates to enrichment effects from
20 nitrogen deposition. For other details on algal nutrition, the reader is referred to reviews by
21 Goldman and Glibert (1982), Button (1985), Kilham and Hecky (1988), and Hecky and
22 Kilham (1988).
23 Denitrification plays a much larger role in nitrogen dynamics in aquatic ecosystems than
24 it does in terrestrial ones. In streams, rivers and lakes, bottom sediments are the main sites
25 for denitrification (see Seitzinger, 1988a) although open-water denitrification has also been
26 reported (Keeney et al., 1971). In lake and stream sediments the main source of NO3",
27 although potentially available from the water column, is NO3" produced when organic matter
28 is broken down within the sediments, and the resulting NH4+ is subsequently oxidized
29 (Seitzinger, 1988a). Denitrification is an especially important process in large rivers and
30 estuaries, and will play a large role in discussions of nitrogen loading to estuaries and near
31 coastal systems (see Section 10.6.4.2). In a recent review of denitrification in freshwater and
August 1991
10-136 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
estuarine systems, Seitzinger (1988a) reported denitrification rates that were 7 to 35% of
nitrogen inputs in large rivers, and 20 to 50% ^of inputs in estuaries. Denitrification in
aquatic ecosystems is an alkalinizing process, consuming one mole of hydrogen for every
mole of NO3" denitrified.
Estimates of denitrification rates range from 54 to 345 fimol • m"2 h"1 in streams with
high rates of organic matter deposition, 12 to 56 /xmol • m"2 h"1 in (nutrient-poor)
oligotrophic lakes, 42 to 171 /imol fimol • m"2 h"1 in eutrophic lakes and 77 to 232 fimol •
m"2 h"1 in estuaries (see Seitzinger, 1988a). These values are in the range where
denitrification can deplete NO3" pools. Rudd et al. (1990) have reported an increase in the
rate of denitrification from less than 0.1 /imol • m"2 h"1 to over 20 fimol • m"2 h"1 in an
oligotrophic lake when nitric acid was added in a whole-lake experimental acidification,
suggesting that freshwater denitrification may limited by NO3" availability. Denitrification
can account for 76 to 100% of nitrogen flux at sediment-water interfaces in rivers, lakes and
estuaries (Seitzinger, 1988a). In the Potomac and Delaware rivers, where organic sediment
deposition is extreme due to sewage inputs, the loss represents 35 and 20%, respectively, of
external nitrogen inputs. In estuaries, it can represent a 50% loss. In deep mud of slow
flowing streams, the process can effectively reduce NO3" concentrations in the water column
by as much as 200 #eq • L"1 over a 2 km length of stream (Kaushik et al., 1975; Chatarpaul
and Robinson, 1979). This depletion amounts to 75% of the daily input of NO3" during a
growing season and it has been sufficient to consider denitrification as a method for NO3"
removal in the management of some slow-moving streams having a deep organic substrate
(Robinson etal., 1979).
Nitrogen fixation counteracts denitrification losses of nitrogen from surface waters and
is fundamental to replenishing fixed forms of nitrogen in all aquatic ecosystems. It is thought
to be the main process responsible for maintaining surplus inorganic nitrogen in lakes and
streams and is fundamental to the fact that primary production in most lakes and streams is
limited by phosphorus (Schindler, 1977). In estuaries, however, there is a higher loss of
nitrogen relative to that fixed or imported. The loss may be due to high rates of
denitrification (Seitzinger, 1988a), which creates relative nitrogen deficiencies.
Rates of nitrogen fixation are generally related to trophic status in freshwater. Howarth
et al. (1988a) show that fixation in low-, medium-, and high-nutrient lakes is generally
August 1991
10-137 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
<0.02, 0.9 to 6.7, and 14.3 to 656.9 mmol • N : m"2 • yr"1 respectively. Fixation is also
closely correlated with the abundance of blue-green algae (Wetzel, 1983), which suggests that
the algae, rather than bacteria, dominate nitrogen fixation in lakes. Although nitrogen
fixation does occur in sediments, that source is of minor importance compared to that in the
water column. Only in very nutrient-poor lakes, where nitrogen loading from all other
sources is small, can nitrogen fixation in sediments gain some significance (e.g., 32% and
6% of total inputs in Lake Tahoe, California, and Mirror Lake, New Hampshire,
respectively; Howarth et al., 1988a).
Unlike the nitrogen fixation community in lakes, nitrogen fixers in estuaries are
dominated by bacteria, producing rates of 0.1 to 111 mmol • N • m"2 • yr"1 (Howarth et al.,
1988a). Highest rates occur in deep organic sediments, but even these are a relatively small
percentage of total nitrogen inputs to estuaries (reviewed by Howarth et al., 1988a).
As in terrestrial watersheds, rates of nitrification in lakes and streams are often limited
by low concentrations of NH4+. Supply rates of NH4+ from watersheds are often low
(except in cases of point-source pollution), and nitrifying organisms have little substrate with
which to work. Two exceptions to this generality are cases where NH4+ deposition is
extremely high, such as near agricultural areas, and cases where NH4+ is produced within
the aquatic system. Experiments on whole lakes and in mesocosms in Canada have
confirmed the acidifying potential of ammonium additions from deposition to surface waters
(Schindler et al., 1985; Schiff and Anderson, 1987). Ammonium deposition is especially
deceptive, because in the atmosphere it can combine as a neutral salt with SO42', resulting in
precipitation with near-neutral pH values, as seen in The Netherlands (van Breemen and van
Dijk, 1988). Once deposited, however, the ammonium can be assimilated, leaving an
equivalent amount of hydrogen, or nitrified, leaving twice the amount of hydrogen. There is
some evidence from Canadian whole-lake experiments that nitrification in lakes is an acid
sensitive process; Rudd et al. (1988) presented data indicating that nitrification was blocked at
pH values less than 5.4 in an experimentally-acidified lake, leading to a progressive
accumulation of NH4+ in the water column.
High NH4+ concentrations may also result in lakes whose deeper waters become anoxic
during periods of stratification (usually late winter or late summer). Production of NH4+ (by
decomposition) can be substantial under anaerobic conditions, and NH4+ may accumulate in
August 1991
10-138 DRAFT-DO NOT QUOTE OR CITE
-------
1 the anoxic water. Nitrification of this NH4+ occurs when lakes mix during spring or fall,
2 supplying the oxygen necessary for nitrifying organisms to survive (Wetzel, 1983). In
3 estuaries, the processes of nitrification (aerobic) and denitrification (anaerobic) may be
4 closely coupled at the sediment surface, with mineralization in the anaerobic sediments
5 supplying NH4+ to nitrifiers at the sediment/water interface (Jenkins and Kemp, 1984).
6 Except in cases where the overlying water becomes anoxic (as may be common in the
7 summer months), the nitrifying organisms supply NO3" back to the sediments for subsequent
8 denitrification. In both cases described above (the annual cycle in lakes, and the
9 sediment/water interface cycle in estuaries), the main influence of nitrification is to recycle
10 nitrogen within the system, and to supply NO3" to either denitrifiers Or to nitrogen-deficient
11 algae.
12 In lakes, streams, and estuaries, water is in constant movement, and to a large extent
13 the effects of nitrogen cycling on biota are regulated by the local hydrology. In lakes,
14 oxidation and reduction reactions are perceived to occur as cycles in the sense that water has
15 a residence time lasting from a few weeks in small ponds to many years in large lakes.
16 Nitrogen species are assimilated, contribute to biological productivity, the organic forms are
17 subsequently mineralized, and the resulting inorganic forms enter various oxidizing and
18 reducing pathways mediated by a microbial community within a single body of water. One
19 or more complete cycles can be followed within a single lake before export downstream.
20 In streams, and to some extent in estuaries, nitrogen dynamics are more closely
21 dependent on the physical movements of water. As nitrogen compounds are cycled among
22 the biotic and abiotic components of the stream ecosystem, they are subject to downstream
23 transport. Among stream ecologists, this coupling between nutrient cycles and water
24 movement is termed "nutrient spiralling" (e.g., Elwood et al., 1980; Newbold et al., 1983).
25 According to this concept, nitrogen cycling occurs in most streams, but little or no recycling
26 occurs in any one place. Nitrogen is instead regenerated or transformed at one point in the
27 stream and transported downstream before subsequent reutilization or retransformation
28 (Stream Solute Workshop, 1990). The movement of water can increase nutrient uptake rates
29 and growth rates in freshwater algae (Whitford and Schumacher, 1961, 1964) by continually
30 resupplying nutrients at cell walls. This constant replenishment prevents steep concentration
31 gradients from becoming established, as can happen in less active lake water (Gavis, 1976).
August 1991
10-139
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
It also maintains high rates of production and nutrient assimilation. Biomass eventually
sloughs from substrata, and drifts as fine particulate organic matter (Meyer and Likens, 1979)
for settlement, decomposition and mineralization downstream. Very high flows associated
with intense precipitation events are physically disruptive and can increase the concentration
of particulates transported downstream (Bilby and Likens, 1979; Holmes et al., 1980).
Efficiencies of nutrient uptake also decrease with increasing flows because of reduced contact
time that a given ion has with the reactive substrate (Meyer, 1979).
One important consequence of nutrient spiralling in streams is that any block in the
nitrogen cycle upstream can have potential effects on nitrogen conditions downstream.
Mulholland et al. (1987), for example, have presented experimental evidence that leaf
decomposition (mineralization) in streams is inhibited at low pH values. Because
mineralization of organic matter is an important process in resupplying nitrogen to organisms
downstream, the existence of acidic headwaters could influence biotic conditions in
downstream portions of streams where acidification is not important.
10.6.3 The Effects of Nitrogen Deposition on Surface Water Acidification
The acidification processes of lakes and streams are conventionally separated into
chronic (long-term) and episodic (event-based) effects. A great deal of emphasis in the past
decade has been placed on chronic acidification in general, and on chronic acidification by
sulfate in particular (e.g., Galloway et al., 1983; Sullivan et al., 1988; Brakke et al., 1989).
This emphasis on sulfate (SO42~) has resulted largely because sulfur deposition rates are often
higher than those for nitrogen (S deposition rates are approximately twice the rates of
N deposition in the Northeast; Stensland et al., 1986), and because NO3" appears to be of
negligible importance in surface waters sampled during summer and fall index periods
(Linthurst et al., 1986). As mentioned previously, summer and fall are seasons when
watershed demand for nitrogen is very high, while supply rates (from deposition) are low,
creating a low probability that nitrogen, in any form, will be leached into soil and surface
waters unless the watersheds have achieved nitrogen saturation. Under conditions of low
nitrogen deposition (or high nitrogen demand), nitrogen leaking from terrestrial ecosystems,
as described earlier, is more likely to be a transient (or seasonal) phenomenon than a chronic
one. As a result, the primary impact of nitrogen in surface water acidification will be
August 1991
10-140 DRAFT-DO NOT QUOTE OR CITE
-------
1 observed during high-flow seasons, and particularly during snow melt. It has been estimated
2 that 40% to 640% more streams in the eastern U.S. are acidic during spring episodes than are
3 acidic during spring base flow, while the number of acidic Adirondack lakes is estimated to
4 be three times higher during the spring than during the fall (Eshleman, 1988).
5 Surface waters are conventionally considered acidic if their acid neutralizing capacity
6 (ANC) is less than zero. The ANC of a lake or stream is a measure of the water's capacity!
7 to buffer acidic inputs, and results from the presence of carbonate and/or bicarbonate (or
8 alkalinity), aluminum and organic acids in the water (Sullivan et al., 1989). The main
9 purposes of this section are to evaluate the evidence for chronic acidification by nitrogen
10 deposition in North America, and to determine what role nitrogen deposition plays in episodic
11 acidification.
12
13 10.6.3.1 Chronic Acidification
14 In the United States, the most comprehensive assessment of chronic acidification of
15 lakes and streams comes from the National Surface Water Survey (NSWS) conducted as part
16 of the National Acid Precipitation Assessment Program (NAPAP). The NSWS surveyed the
17 acid/base chemistry of both lakes and streams using an "index period" concept. The goal of
18 the index period concept was to identify a single season of the year that exhibited low
19 temporal and spatial variability and that, when sampled, would allow the general condition of
20 surface waters to be assessed (Linthurst et al., 1986). In the case of lakes, the index period
21 selected was autumn overturn (the period when most lakes are mixed uniformly from top to
22 bottom), while in streams the chosen index period was spring base flow (the period after
23 spring snow melt and before leaf-out) (Messer et al., 1988). Because of the strong
24 seasonality of the nitrogen cycle in forested watersheds (described earlier) the choice of index
25 period plays a very large role in the assessment of whether nitrogen is an important
26 component of acidification.
27 The results of the Eastern Lake Survey (Linthurst et al, 1986), based on a.probability
28 sampling of lakes during fall overturn, suggest that nitrogen compounds make only a small
29 contribution to chronic acidification in North America. Henriksen (1988) has proposed that
30 the ratio of NO3~:NO3~+SO42~ in surface waters be used as an index of the influence of
31 NO3" on chronic acidification status. This index assesses the importance of nitrogen relative
August 1991
10-141
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
to the importance of SO42~, which is usually considered more important in chronic
acidification (see above). A value greater than 0.5 indicates that NO3" has a greater influence
on the chronic acid/base status of surface waters than does SO42~. Henriksen (1988)
summarized the ratios for acid-sensitive sites worldwide; .these results are repeated in
Table 10-21. In general, Henriksen's results show that NO3" can be as important as SO42" in
some parts of Europe, but that ratios are low in the United States, except for selected
Adirondack systems (see also Henriksen and Brakke, 1988).
One problem with Henriksen's approach, however, is that he compares data collected
intensively (i.e., through multiple samples per year) with survey data collected during a
single index period. The data presented for Adirondack lakes in Table 10-21, for example,
were collected monthly over a two year period (Driscoll and Newton, 1985), and the
apparent difference between the Adirondacks and the rest of central New England (from the
regional survey data) could well result from comparing fall values to annual mean values,
Annual mean values include high spring NO3" concentrations and will be therefore be higher
than concentrations measured only in the autumn. As a result, the high ratio values reported
in Table 10-21 for the Adirondacks are a good indication that NO3" may be important in
chronic acidification (i.e., NO3" makes up about 15% of acid anions), but the low ratios
reported for the Eastern Lake Survey are not informative. Unfortunately, no regional lake
survey with representative annual, or spring, values exists for the United States, and
questions concerning the role of NO3" in chronic lake acidification remain unanswered for
areas outside of the Adirondacks.
Values of NO3~:NO3~+SO42" ratios are also available for streams from the National
Stream Survey (NSS; Kaufmann et al., 1988), as well as from other regional stream surveys
(e.g., Stoddard and Murdoch, 1991). Median values for each of the regions covered in these
surveys are given in Table 10-22. The NSS data have the advantage of having been collected
during a spring baseflow index period. This period is been shown to be a good index of
mean annual condition for streams (Messer et al., 1988; Kaufmann et al., 1988), but is not
an estimate of worst case condition, as concentrations taken during spring snow melt would
be. The Catskill regional data included in Table 10-22 are from a stream survey which
included multiple samplings per year (Stoddard and Murdoch, 1991). Several stream regions
exhibit ratios as high as those reported for the Adirondacks by Henriksen (1988). Several
August 1991
10-142
DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-21. CONCENTRATIONS OF NITRATE, SULFATE, AND RATIOS
OF NITRATE TO THE SUM OF NITRATE AND SULFATE IN RUNOFF WATERS
IN ACIDIFIED AREAS OF THE WORLD (FROM HENREKSEN, 1988)
Location
West Germany
Lange Bramke
Lange Bramke
Baverischer Wald
Rachelsee
Gr. Arbersee
Kl. Arbersee
Poland
'The Giant Mts.
Maly Staw
Wielki Staw
Czechoslovakia
Tatra Mts.
Av 53 lakes
Jameke
Popradake
:Vyshe Wahlenbugoro
Vyshe Furkotake
Bohemia
Carne
Certovo
Prasilske
Plesne
Laka (man-made)
Zdarske (man-made)
Krusne hory Mts.
Sumava Mts.
Liz
Albrechtec
Norway
Birkenes
Storgama
Sweden
Stromyra
Scotland
Av 22 lakes in
the Galloway area
Year
1977
1984
1985
• 1985
1985
1986
1986
1984
1980-82
1980-82
1980-82
1980-82
1986
1986
1986
19.86
1986
1986
1986
Apr 86
Apr 86
1973-86
1973-86
1984-85
1979 '
PH
5.8
. 6.2
4.5
4.7
4.5
5.5
4.7
6.1
4.4
6.6
5.6 „
6.3
4.5
4.2
4.5
4.7
5.5
6.5
5.2
5.89
6.22
4.52
4.56
6,54
4.97
A*eq
N03-
16
49
77
9.8
93
13
40
37
2
40
44
42
93 .
85
40
41
45
0
118
136
36
9
12
17
21
•L-1
so/-
233
230
135
118
108
92
140'
97
171
111
74
110
152
182
120
203
61
156
1216
390
358
140
77
- 180
103
NO3-:N03- + SO/'
0.06
,0.18
0.36
0.45
0.46
0.12
0.22
0.27
0.01
0.26
0.37 ,
0.28
0.38
0.32 ,
0.25
0.17
0.42
0.00
0.09
0.26 ,
0.09
0.06
0.13
•
0.09
0.17
Sampling
Methoda
Intensive
Intensive
Unknown
Unknown
Unknown
Unknown
Unknown
Unknown
Unknown
Unknown
•Unknown
Unknown
Unknown
Unknown
Unknown
Unknown
Unknown
Unknown
Unknown
. Unknown
Unknown
Intensive
Intensive
Intensive
E '. - " -
Unknown
August 1991
10-143 DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-21 (eont'd). CONCENTRATIONS OF NITRATE, SULFATE, AND
RATIOS OF NITRATE TO THE SUM OF NITRATE AND SULFATE IN RUNOFF
WATERS IN ACIDIFIED AREAS OF THE WORLD (FROM HENREKSEN, 1988)
Location
Year
pH
SO/
NO3-:NO3-
Sampling
Method3
United States
Adirondacks
Big Moose Lake
Cascade Lake
Darts Lake
Merriam Lake
Lake Rondaxe
Squash Pond
Townsend Pond
Windfall Pond
Bubb Lake
Constable Pond
Moss Lake
Black Pond
Clear Pond
Heart Lake
Otter Lake
West Pond
Woodruff Pond
Eastern Lake Survey**
Southern Blue Ridge
Florida
Upper Midwest
Upper Great Lakes
Wisconsin
Peninsula, Michigan
NE Minnesota
Maine
S New England
C New England
Canada
1980s
1985
5.1
6.5
5.2
6.4
5.9
4.6
5.2
5.9
6.1
5.2
6.4
6.8
7.0
6.4
5.5
5.2
6.9
24
29
24
26
23
24
27
26
16
17
26
4
1
5
9
10
2
3
1
0.7
0.6
1.0
0.6
0.9
0.2
0.8
0.3
140
139
139
141
134
131
154
141
131
149
141
130
139
106
138
111
147
32
94
57
50
57
78
62
75
141
101
0.15
0.17
0.15
0.16
0.15
0.15
0.15
0.16
0.11
0.10
0.16
0.03
0.00
0.05
0.06
0.08
0.01
0.09
0.01
0.01
0.01
0.02
0.01
0.01
0.00
0.01
0.00
Monthly
Monthly
Monthly
Monthly
Monthly
Monthly
Monthly
Monthly
Monthly
Monthly
. Monthly
Monthly
Monthly
Monthly
Monthly
Monthly
Monthly
Fall index
Fall index
Fall index
Fall index
Fall index
Fall index
Fall index
Fall index
Fall index
Fall index
Experimental Lakes
Area, Ontario
Sudbury, Ontario
Kekimkujik,
Nova Scotia
1980s
1980s
1980s
1
2
2
3
78
252
252
78
0.01
0.01
0.01
0.04
Intensive
Intensive
Intensive
"Sampling methods are listed as either monthly, intensive (more frequent than monthly) or based on a single fall
index sample.
"Median value for regional population of lakes.
August 1991
10-144 DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-22. CONCENTRATIONS OF NITRATE, SULFATE, AND RATIOS
OF NITRATE TO THE SUM OF NITRATE AND SULFATE IN STREAMS OF
ACID-SENSITIVE REGIONS OF THE UNITED STATES. VALUES ARE
MEDIANS FOR REGION (FIRST AND THIRD QUARTBLES IN PARENTHESES)
Location
Poconos/Catskills
Northern Appalachians
Valley and Ridge
Mid-Atlantic Coastal Plainb
Southern Blue Ridge
Piedmont
Southern Appalachians
Ozarks/Ouachitas
Florida
Catskill Regional Survey0
Median value for 51 streams
Year pH
National Stream Survey*
1986 6.96
6.60
7.05
5-98
6.99
6.80
7.33
6.62
5.48
1984-86 6.60
/*e;
N03-
6
(2-18)
30
(12-4
1)'
10
(3-31)
-
8
(2-16)
2
(0-5)
16
(3-32)
1
(1-4)
5
(1-10)
29
(14-4
7)
g-L'1
SO/
169
(154-184)
171
(135-347)
154
(84-294)
-
17
(10-27)
48 ,
(19-63)
58
(30-104)
59
(48-83)
22
(9-30)
138
(125-151)
N03':N03- + SO/
0.03
(0.01-0.10)
0.14 •
(0.02-0.19)
0.09
(0.01-0.22)
-
0.28
(0.08-0.44)
0.03
(0-0.20)
0.32
, ' . (0.04-0.40)
0.02 •
(0-0.06)
-'.".. . •
0,19
. (0.10-0.25)
0.17 ;
•. ; (0.09-0.26)
"Values for pH are for entire region (Kaufmann et al., 1988); medians for NO3", SO/ and the
NO3':NO3" + SO/ ratio exclude sites with potential agricultural or other land use impacts (Baker et al.,
in press).
bThe influence of agricultural and land use practices could not be ruled out for any of the sites in the
Mid-Atlantic Coastal Plain (Baker etal., in press). ,
Trom Stoddard and Murdoch (1991).
1
2
3
regions in the southeastern United States exhibit high ratios in part because their current
SO42" concentrations are relatively low., The Southern Blue Ridge, in particular, has the
lowest NO3" concentrations found in the NSS, and the relatively high NO3":NO3"+SO42"
August 1991
10-145 ' DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
ratios in this region could be considered misleading. The stream data do suggest that the
Catsldlls, Northern Appalachians, Valley and Ridge Province, and Southern Appalachians all
show some potential for chronic acidification due to NO3". In all of the stream regions in
Table 10-22, as well as the lake regions in Table 10-21, however, chronic acidification is
more closely tied to SO42" than to NO3".
The data presented thus far in this section establish which regions of the country show
potential problems with chronic acidification by NO3", but do not indicate whether the source
of the NO3" is atmospheric deposition. As described earlier, several watershed processes
(e.g., mineralization, nitrification, and nitrogen fixation) may combine to produce NO3", and
may be responsible, at least in part, for high NO3" concentrations observed in surface waters.
On a regional scale, it is not possible to attribute surface water NO3" to any single source,
but two efforts have been made to relate rates of nitrogen deposition to rates of nitrogen loss
from watersheds. Data from the NSS (Kaufmamret al., 1991) suggest a strong correlation
between concentrations of streamwater nitrogen (NO3~ + NH4+) at spring base flow and
levels of wet nitrogen deposition (NO3~ + NH4+) in each of the NSS regions
(Figure 10-25a). The only exception to this relationship is the Pocono/Catskill region, where
nitrogen deposition is highest (450 eq • ha"1 • yr"1), but where stream water nitrogen
concentrations fall below what is expected, based on results from the other regions. The
median streamwater NO3" value for the Catskills alone (from Stoddard and Murdoch, 1991;
Table 10-22) is 29 /ieq-L"1, and fits the relationship much more closely, suggesting that
watersheds in the southern portion of this region (the Poconos) are retaining nitrogen more
strongly than the northern portion. Driscoll et al. (1989) collected input/output budget data
for a large number of watersheds in the United States and Canada, and summarized the
relationship between nitrogen export and nitrogen deposition at all of the sites
(Figure 10-25b). The authors stress that the data illustrated intFigure 10-25b were collected
using widely differing methods and over various time scales (from one year to several
decades). Given the numerous possible sources of NO3~, and the, watershed pathways
through which nitrogen may be cycled, the relationships illustrated in Figure 10-25 should not
be over-interpreted, nor should they be construed as an illustration of cause and effect.
However, the relationships do show that watersheds in mariy regions of North America are
retaining less than 75% of the nitrogen that enters them, and that the amount of nitrogen
August 1991
10-146 . DRAFT-DO NOT QUOTE OR CITE
-------
50 +
(a) NSS-1 Subregions
100 200 300 400 500
Wet NO" + NH4+Deposition (eq/ha/yr)
o>
at
DC
•e
O)
O
4UU -r
350 -
300-
250-
200-
150-
100 _
50 _
0
(b) o
0 0
O o
-.:••- 0 • °
o
o
o •
o
0 0 0
8 ° • °
o: o
<-» O O O
-> o o Q.—SQ ,— — n i ° °i Q 1
100 200 300 400 500 600
Rate of Nitrogen Wet Deposition (eq/ha" 1/yr"1)
Figure 10-25. (a) Relationship between median wet deposition of nitrogen (NO3" +
NH4+) and median surface water nitrogen (NO3~ + NH4 )
concentrations, for physiographic districts within the National Stream
Survey that have minimal agricultural activity. [Subregions are:
Poconos/Catskills (ID), Southern Blue Ridge Province (2As), Valley and
Ridge Province (2Bn), Northern Appalachians (2Cn), Ozarks/Ouachitas
(2D), Southern Appalachians (2X), Piedmont (3A), Mid-Atlantic Coastal
Plain (3B), and Florida (3C)]. From Baker et al. (in press), (b)
Relationship between wet deposition of nitrogen (NO3~ + NH4 ) and rate
of nitrogen export for watershed studies throughout North America.
Sites with significant internal sources of nitrogen (e.g., from alder trees)
have been excluded. Data from Driscoll and Schaefer (1989).
August 1991
10-147 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
being leaked from these watersheds is higher in areas where nitrogen deposition is highest.
This pattern is consistent with what we would expect if large areas of the eastern United
States were experiencing the early stages of nitrogen saturation. Furthermore, both analyses
suggest a threshold value of nitrogen deposition (between 200 and 300 eq • ha"1 • yr"1) above
which substantial watershed losses of nitrogen might begin to occur.
Chronic acidification due to nitrogen deposition is much more common in Europe than
in North America (Hauhs et al., 1989). Many sites show chronic increases in nitrogen export
from their watersheds (e.g., Henriksen and Brakke, 1988; Hauhs, 1989), and at sites with the
highest stream water NO3" concentrations (i.e., Lange Bramke and Dicke Bramke in West
Germany) NO3" concentrations no longer show the seasonally which is expected from normal
watershed processes (Hauhs et al., 1989). Henriksen and Brakke (1988) have reported
regional chronic increases in surface water NO3" in Scandinavia in the past decade. These
increases in NO3"concentration are associated with increasing concentrations of aluminum,
which is toxic to many fish species (Henriksen et al, 1988; Brown, 1988). There is some
evidence that NO3" has a greater ability to mobilize toxic aluminum from soils than does
SO4 " (James and Riha, 1989). Chronic acidification attributable to ammonium deposition
has also been demonstrated in The Netherlands (van Breemen and van Dijk, 1988; Schuurkes,
1986, 1987). As described earlier, ammonium in deposition can be nitrified to produce both
NO3" and hydrogen ions, which are subsequently leaked into surface waters. Rates of NO3"
and NH4+ deposition are much higher in Europe (in some places deposition is > 2>000 eq •
ha"1 • yr"1; Rosen, 1988) than in the United States (Table 10-20), and it has been suggested
that chronic nitrogen acidification is more evident in Europe than in North America because
nitrogen saturation (see discussion above) is further progressed in Europe.
10.6.3.2 Episodic Acidification
In a recent comprehensive examination, Wigington et al. (1989) reported that acidic
episodes have now been observed in a wide range of geographic locations in Scandinavia
(Norway, Sweden, Finland), Europe (United Kingdom, Scotland, Federal Republic of
Germany, Czechoslovakia), and Canada (Ontario, Quebec, Nova Scotia), as well as the
United States. In the United States, they noted that episodes have been registered in surface
waters in the Northeast, Mid-Atlantic, Mid-Atlantic Coastal Plain, Southeast, Upper
August 1991
10-148 DRAFT-DO NOT QUOTE OR CITE
-------
1 Midwest, and West regions. In the Mid-Atlantic Coastal Plain and Southeast regions, all of
2 the episodes cataloged to date have been associated with rainfall. In contrast, most of the
3 episodes in the other regions are related to snow melt, although rain-driven episodes
4 apparently can occur in all regions of the country.
5 The regional importance and severity of episodic acidification have not been quantified;
6: that is, the regional information on chronic acidification that was gained from the NSWS has
7 no parallel in episodic acidification. As a result, all of the information we currently have
• 8 about the importance of episodes, and the influence of nitrogen deposition on episodes, comes
9 from site-specific studies. It is important to stress that even within a given area, such as the
10 Northeast, major differences can be evident in the occurrence, nature, location (lakes or
11 streams), and timing of episodes at different sites.
12 Eshleman (1988) has used a simple stream mixing model (Johnson et al., 1969) to
13 predict the number of streams in the NSS that would be acidic during spring episodes, based
14 on their spring baseflow chemistry. In addition, Eshleman used an empirical model relating
15 fall index period lake chemistry to spring episodic chemistry, using data from EPA's
16 Long-Term Monitoring (LTM) project (Newell et al., 1987), to predict the number of
17 Adirondack lakes that undergo episodic acidification. His results are repeated in Table 10-23.
18 Eshleman's approach has been criticized (see discussion below), largely because it assumes
19 that all lakes, regardless of their baseline ANC, undergo the same relative depression in ANC
20 during episodes (i.e., that the relationship between fall and spring ANC is linear). This
21 assumption ignores any effect of increased NO3" during episodes, which may be greater in
22 low ANC lakes (Schaefer et al., 1990; Schaefer and Driscoll, in press). Given this criticism,
23 Eshleman's estimates of the number of episodically acidified systems should probably be
24 considered conservative.
25 A number of processes contribute to the timing and severity of acidic episodes (Driscoll
26 and Schaefer, 1989). The most important of these processes are:
27 • dilution of base cations (Galloway et al., 1980) by high discharge;
28
29 • increases in organic acid concentrations (Sullivan et al., 1986) during periods of high
30 discharge;
31
32 • increases in SO42" concentrations (Johannessen et al., 1980) during periods of high
33 discharge; and
August 1991
10-149 DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-23. ESTIMATES OF THE NUMBER AND PROPORTION OF
CHRONICALLY AND EPISODICALLY ACIDIC LAKES AND STREAM REACHES
IN THE EASTERN UNITED STATES. CHRONIC CONDITIONS BASED ON
RANDOM SAMPLE OF SYSTEMS DURING INDEX CONDITIONS (SPRING
BASEFLOW OR FALL OVERTURN). EPISODIC CONDITIONS ESTIMATED
FROM TWO-BOX MIXING MODEL (FOR STREAMS), OR EMPIRICAL
RELATIONSHIPS BETWEEN FALL INDEX PERIOD AND SPRING SNOW MELT
CHEMISTRY (FOR LAKES) (FROM ESHLEMAN, 1988)
Index Conditions (ANC <0),
Subregion*
Stream subregions
Poconos/Catskills
Southern Blue Ridge
Valley and Ridge
Northern Appalachian Plateau
Ozarks/Ouachitas
Southern Appalachians
Piedmont
Mid-Atlantic Coastal Plain
Florida
Lake subregions
Adirondacks
Number
209
0
636
499
0
121
0
1,334
678
138
Proportion (%)
6.4
0
4.9
5.8
0
2.5
0
11.8
39.2
10.7
Episodic Conditions (ANC <0)
Number
746
39
1,126
3,224
75
364
0
3,132
963
459
Proportion (%)
23.0
2.2
8.6
37.2
1.8
7.4
0
27.8
55.7
35.6
'For streams, all data are from the upper end of sampled stream reaches (Kaufmann et al., 1988), except for the
Southern Blue Ridge, where data from lower ends of stream reaches were used.
1
2
3
4
5
6
7
8
9
• increases in NO3" concentrations (Galloway et al., 1980; Driscoll and Schafran,
1984; Schofield et al., 1985) during periods of high discharge.
In addition to these factors, which produce the chemical conditions characteristic of episodic
events, the likelihood of an acidic episode is also influenced by the chemical conditions
before the episode begins. Episodes are more likely to be acidic, for example, if the
baseflow ANC of the stream or lake is low. In this way, acid anions, especially SO42", can
contribute to the severity of an acidic episode, even though they do not increase during the
event, by lowering the baseflow ANC of the stream or lake (Stoddard and Murdoch, 1991).
August 1991
10-150 DRAFT-DO NOT QUOTE OR CITE
-------
1 In many cases, all of these processes will contribute to episodes in a single aquatic
2 system. Dilution, for example, probably plays a role in all episodic decreases in ANC and
3 pH in all regions of the United States (Wigington et al., 1989). Dilution results from the
4 increased rate of runoff, and channeling of runoff through shallower soil layers, that occurs
5 during storms or snow melt; the shorter contact time produces runoff with a chemical
6 composition closer to that of atmospheric deposition than is typical of baseflow conditions
7 (e.g., Driscoll and Newton, 1985; Peters and Murdoch, 1985; Stoddard, 1987a). Because
8 precipitation is usually more dilute than stream or lake water, storm runoff produces surface
9 waters that are more dilute than during non-runoff periods. In a sense, dilution sets the
10 baseline condition to which is added the effects of organic acids and atmospherically derived
11 SO42'and NO3'.
12 Little information exists about the effects of changes in organic acids during episodes.
13 Driscoll et al. (1987a) and Eshleman and Hemond (1985) concluded that organic acids did
14 not contribute to snowmelt episodes in the Adirondacks or in Massachusetts, respectively. At
15 Harp Lake in Canada, organic acidity is believed to remain constant (Servos and Mackie,
16 1986) or decrease (LaZerte and Dillon, 1984) during snowmelt episodes. Haines (1987) and
17 McAvoy (1989) have documented increases in organic acidity during rain-caused episodes in
18 coastal Maine and in Massachusetts.
19 Storage of SO42~ in watersheds, and subsequent release of SO42" during episodic events,
20 is well documented in many parts of Europe (Wigington et al., 1989), but has not been
21 commonly found in the United States. Sulfate episodes have been described for the Leading
22 Ridge area of Pennsylvania (Lynch et al., 1986) and at Filsen Creek in Minnesota (Schnoor
23 et al., 1984), but are not widespread. Sulfate does contribute to episodic acidity, however, in
24 the sense that concentrations may remain high during events, and contribute to a lower
25 baseline ANC; the effects of other factors, such as increased NO3", will be in addition to any
26 constant effect of SO42~ in lowering the baseline ANC (Stoddard and Murdoch, 1991). ,
27 The main goal of this section is to determine when increases in NO3" concentrations
28 play a significant role in episodic acidification. In the Adirondacks, for example, strong
29 NO3" pulses in both lakes (Galloway et al., 1980; Driscoll and Schafran, 1984) and streams
30 (Driscoll et al., 1987b) are apparently the primary factor contributing to depressed ANC and
31 pH during snow melt. Schaefer et al. (1990) examined the same empirical relationships used
August 1991
10-151 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
for the Adirondack lakes by Eshleman (1988; Table 10-23), and conclude that the magnitude
of the episodes experienced by lakes depends strongly on their base cation concentration.
They conclude that lakes with high base cation concentrations (and therefore high ANC
values) undergo episodes that are largely the result of dilution by snow melt. Low ANC
lakes, on the other hand, undergo episodes that result largely from increases in NO3"
concentrations. At intermediate ANC levels, lakes are affected by both base cation dilution,
and NO3" increases, and therefore these lakes may undergo the greatest increases in acidity
during snowmelt episodes (Figure 10-26). The relationship between spring and fall lake
chemistry is therefore not linear, as assumed by Eshleman (1988), and the number of lakes
that become acidic during spring episodes is probably larger than predicted in Table 10-23.
Driscoll et al. (1989a, 1989b) report on a detailed study of nitrogen dynamics in
Pancake-Hall Creek in the Adirondack Mountains. This stream is highly acidic, with low
and invariant concentrations of base cations, and high and invariant concentrations of SO42"
(Figure 10-27). Nitrate concentrations were lower than SO42", and exhibited a distinct
seasonal pattern; peak concentrations approached 100 /zeq • L"1. Short-term changes in NO3"
were highly correlated, and chemically consistent, with changes in the concentrations of
acidic cations (hydrogen and aluminum) (Driscoll et al., 1989a). As mentioned earlier, while
dilution of base cations and increases in NO3" appear to be the primary causes of episodic
acidification in Pancake-Hall Creek, these episodes are excursions from an already low
baseline ANC, which can be largely attributed to high SO42' concentrations.
Stoddard and Murdoch (1991) have concluded that increases in NO3", base cation
dilution, and high baseline SO42" concentrations all contribute to acidic episodes in Catskill
Mountain streams (Figure 10-28). In Biscuit Brook, an intensively-studied stream in the
CatsktUs, concentrations of NO3" approach those of SO42" during episodes (Murdoch and
Stoddard, in review). Values for the ratio of NO3" to NO3" + SO42", as presented in
Tables 10-21 and 10-22, illustrate both the general importance of NO3" to the acid/base
dynamics of this stream, and the increase in importance of NO3" during high-flow events
(Figure 10-28).
Researchers at the Hubbard Brook Experimental Forest in New Hampshire have been
studying the links between atmospheric deposition, watershed processes, and stream water
chemistry since 1963 (Likens et al., 1977). In reference Watershed #6, stream water NO3"
August 1991
10-152 DRAFT-DO NOT QUOTE OR CITE
-------
v 120l
*J
If 100-
1 80-
«p 60-
J>
cL 20'
§2 °
•s -<
•
•
• 1
i
W ' o 40 80 120 160 200 240
Baseline ANC; (jieqL" )
g ^
^ £.' 1
^>
-------
, VCSQ^T, ,^T
1 J~ 1 1—•—
M S J
1985
THWE(monlhs)
1996
Figure 10-27. Temporal patterns in the chemical characteristics of stream water at
Pancake-Hall Creek in the Adirondacks. Sulfate and base cation
concentrations are relatively invariant, while NO3" concentrations
undergo strong seasonally driven by snow melt. Increases in inorganic
monomeric aluminum result when ANC values fall below zero. From
Driscoll et al. (1989a).
August 1991
10-154 DRAFT-DO NOT QUOTE OR CITE
-------
100
^ o
1 -5
* £
"£ I
*=• s
0.0
9
6
7
6
S
4
3
2
I
0
I l t I •! I I .I—It II I ' I I
•+—t-
J FUAUJJA.SONOJ FMAUJJASONO
1988 1989
Figure 10-28. Temporal patterns in chemical characteristics of stream water at Biscuit
Brook in the CatskiU Mountains. AU chemical variables undergo strong
seasonally, with strong dependence on stream discharge. Values for the
ratio of NO3" to NO3" + SO42" approach 0.5 during episodes, and
indicate that NO3" is nearly as important an acidifying influence as SO4 "
during high flow events. Data from Murdoch and Stoddard (in review).
August 1991
10-155 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
concentrations undergo strong seasonal cycles, with peak concentrations as high as
85 /zeq • L" . Both NO3" and hydrogen ion concentrations increase during snow melt at
Hubbard Brook, while SO42' concentrations decrease slightly (Johnson et al., 1981; Likens,
1985). -..-.'',
The highest recorded NO3" concentrations in streams draining undisturbed watersheds in
the United States come from the Great Smoky Mountains in Tennessee and North Carolina.
Nitrate concentrations in Raven Fork (Jones et al., 1983), Clingman's Creek, and Cosby
Creek (Elwood et al., 1991) range from 50 to 100 /ieq • I/1, and in all cases are comparable
to, or higher than, SO42" concentrations. In a survey of stream chemistry at a large number
of sites in the Smokies, Silsbee and Larson (1982) reported NO3" concentrations ranging from
0.2 to 90 #eq . L'1; NO3~ concentrations were highest at higher elevations, and in areas of
old-growth spruce-fir forest that have never been logged. In many cases, NO3"
concentrations in streams of the Smoky Mountains are higher than nitrogen concentrations in
deposition, suggesting both that rates of biological nitrogen uptake are low, and that
mineralization rates are high (Joslin et al., 1987). Unfortunately, few data are available to
suggest the original source of nitrogen now being mineralized in this region. Unless nitrogen
fixation rates have been historically quite high, at least some of the NO3~ now being leaked
from watersheds in the Smokies must have originated as atmospheric deposition. The data of
Silsbee and Larson (1982) suggest strongly that forest maturation is linked to the process of
NO3' leakage from Great Smoky Mountain watersheds; mineralization of soil nitrogen
appears to be high only in old-growth forests (Elwood et al., 1991).
In Canada, the influence of NO3~ on episodic acidification is less universal. Molot
et al. (1989) and DriscoU et al. (1989a) report on numerous episodic events in 15 streams in
the Harp, Dickie and Plastic lake watersheds. Most of these events were driven by base
cation dilution; only one event was dominated by increases in NO3" concentration. The
authors conclude that NO3- plays at least a small role in most episodes, and that NO3"
increases play a greater role in acidic systems than in non-acidic ones.
Small increases in NO3~ concentrations during hydrologic events have been recorded at
sites in a few remaining areas of North America, including northeastern Georgia (Buell and
Peters, 1988), where maximum concentrations were ca. 12 /jeq • L"1. Several studies have
reported the existence of NO3" episodes in the western United States, including the North
August 1991
10-156 DRAFT-DO NOT QUOTE OR CITE
-------
1 Cascades (Loranger and Brakke, 1988), and Sierra Nevada Mountains (Melack and Stoddard,
2 1991). In general, the maximum concentrations of NO3" observed in the West are less than
3 15 Meq • L"1, substantially lower than in most of the eastern United States. Lakes in the
4 mountainous West, however, tend to be much more dilute, and therefore more sensitive to
5; ,:. acidic deposition than in the East. Thirty-nine percent of lakes in the Sierra Nevada, for
6 .example, have ANG values less than 50 jaeq • L"1, as do 26% of the lakes in the Oregon
7 Cascades, and 17% of the lakes in the North Cascades (Landers et al., 1987). Combined
8 , with base cation dilution and small concentrations of SO42", the NO3" increases observed
9 during episodes at Emerald Lake, in the Sierra Nevada, have been sufficient to drive ANC to
10 . zero on two occasions in the past 4 years (Williams and Melack, in press b). Data from the
11 outflow at Emerald Lake in 1986 and 1987 (Figure 10-29) indicate that minimum ANC
12 values are coincident with maximum concentrations of NO3" and diluted base cation
13 , concentrations. It should be noted, however, that at no time has the pH of Emerald Lake
14 fallen below 5.5, a level commonly considered the threshold for injury to fish populations,
15 and that ANC values of zero can be caused by base cation dilution alone (a natural process).
16 The state of episodic acidification in the Sierra (and the rest of the West) remains therefore
17 uncertain, because few data exist and the data that are available indicate ANC depressions to
18 , a value of 0-#eq • L"1, but not below.
19 Finally, there are some areas of North America where no significant affect of NO3~ on
20 episodic acidification has been observed. Morgan and Good (1988) report data on 10 streams
21 in the New Jersey Pine Barrens, and found mean annual NO3~ greater than 1 /aeq • L" only
22 in disturbed streams (in residential and agricultural watersheds). Swistock et al. (1989) and
23 Sharpe et al. (1984, 1987, 1989) reported data on episodic acidification of several streams in
24 the Laurel Hill area of southwestern Pennsylvania and found that NO3" played only a minor
25 role in stream acidification and fish kills. Baird et al. (1987) examined episodic acidification
26 during snow melt at Cone Pond, New Hampshire, and were unable to detect any NO3~ in
27 inlet water. Cosby et al. (1991) have examined 7 years of data from two streams in Virginia,
28 and found no evidence of NO3" episodes; NO3" concentrations are always less than 15 /zeq •
29 L"1 in these streams. Swank and Waide (1988) reported data from 7 undisturbed watersheds
30 , . at the Coweeta Hydrologic Laboratory in North Carolina, where ,the volume-weighted mean
31
concentrations of NO3" were less than 1.5 jueq • L
-1
August 1991
10-157 DRAFT-DO NOT QUOTE OR CITE
-------
Emerald Lake, Sierra Nevada, California
cr
CD
a e-
^ 6-
c
£D
O
cr
o
O
9
6<
3-
0-
i i i i t I
cr
1986
hw'A'M'j
1987
2-
1.5-
0.5-
•
0
( I (IT I
60-
45-
30-
'F'M'A'M'J
1986
i i r
1987
Figure 10-29. Outflow chemistry from two snowmelt seasons (1986 and 1987) at
Emerald Lake, a high elevation lake in the Sierra Nevada mountains of
California. Nitrate episodes are smaller in magnitude than at sites in the
eastern United States, but western lakes may be more susceptible to
episodic acidification because of their lower baseline acid neutralizing
capacity than most eastern lakes.
August 1991
10-158 DRAFT-DO NOT QUOTE OR CITE
-------
1 Some broad geographic patterns in the frequency of episodes in the United States are
2 now evident. Acidic episodes driven by NO3~ are apparently common in the Adirondack and
3 Catskill Mountains of New York, especially during snow melt, and also occur in at least
4 some streams in other portions of the Northeast (e.g., at Hubbard Brook). Nitrate contributes
5 on a smaller scale to episodes in Ontario, and may play some role in episodic acidification in
6 the western United States. There is little current evidence that NO3" episodes are important
7 in the acid-sensitive portions of the southeastern United States outside of the Great Smoky
8 Mountains., We have no information on the importance of NO3~ in driving episodes in many
9 of the subregions covered by the NSS, including those that exhibited elevated NO3"
10 concentrations at spring baseflow (e.g., the Valley and Ridge Province and Mid-Atlantic
11 Coastal Plain), because temporally-intensive studies have not been published for these areas.
12 As was the case with chronic acidification discussed earlier, the mere presence of NO3~
13 in acidic episodes should not be construed as proof that nitrogen deposition is having an
14 acidifying effect on surface waters; many other sources of nitrogen exist in watersheds.
15 There is currently little direct evidence linking nitrogen deposition with those acidic episodes
16 that are driven by increases in NO3" concentrations, at least partially because the type of data
17 necessary to link deposition to stream water pulses of NO3~ are extremely difficult to collect.
18 High concentrations of NO3~ during snow melt may simply result when NO3" stored in the
19 snow pack during the winter months is released while the forest is still dormant. The reduced
20 biological activity typical of the winter months creates less demand for nitrogen, and
21 snowpack NO3" may simply runoff without entering the nitrogen, cycle of the forest or
22 watershed. Several mechanisms, however, will amplify the signal produced by atmospheric
23 deposition of nitrogen to snow packs. In areas with large snow packs (e.g., much of the
24 Northeast and all of the mountainous West) ions have been shown to drain from the pack in
25 the early stages of snow melt, leading to concentrations that are much higher than the average
26 concentration of the snowpack itself (e.g., Jeffries, 1990). This differential elution of acid
27 anions (like NO3") during the'initial stages of snow melt has been shown to be responsible for
28 the elevated NO3" concentrations observed in parts of Scandinavia (Johannessen and
29 Henriksen, 1978), Canada (Jeffries^ 1990), the Adirondacks (Mollitor and Raynal, 1982), the
30 Midwest (Cadle et al., 1984), and in the Sierra Nevada (Williams and Melack, in press b).
31 Ammonium deposited to the snowpack as either wet or dry deposition can be subsequently
• August 1991
10-159 DRAFT-DO NOT QUOTE OR CITE
-------
1 nitrified to NO3" in soils, or while still in the snowpack, and produce NO3" concentrations
2 elevated over those calculated from NO3" deposition alone (Galloway et a!., 1980; Schofield
3 et al., 1985; Cadle et al., 1987; Schaefer and Driscoll, in press). Rates of dry deposition of
4 nitrogen compounds to the snowpack are difficult to measure, but potentially important,
5 controls on NO3" concentrations in snowmelt water (Galloway et al., 1980; Cadle et al.,
6 1987). Jeffries (1990) presents a recent review of snowpack storage and release of pollutants
7 during snow melt.
8 Some evidence does exist that mechanisms other than atmospheric deposition contribute
9 to NO3" episodes, at least on a small scale. Rascher et al. (1987), for example, have shown
10 that mineralization of organic matter in the soil during the winter months, and subsequent
11 • nitrification, contribute substantially to snowmelt NO3" concentrations at one site in the
12 Adirondacks. Schaefer and Driscoll (in press) have suggested that a similar phenomenon
13 contributes to NO3" pulses during snow melt at 11 Adirondack lakes, and that the
14 contribution from mineralization is greater in low ANC and acidic lakes. Stottlemyer and
15 ToczydlowsM (1990) also report that mineralization contributes to snowmelt NO3" at a site on
16 the upper peninsula of Michigan. It is not currently known how widespread this phenomenon
17 is. Murdoch and Stoddard (in review) conclude that mineralization probably does not
18 contribute substantially to NO3" episodes in the Catskill Mountains, because maximum NO3"
19 concentrations are very similar among a large number of streams; mineralization rates are
20 expected to differ among watersheds, and would produce variability in concentrations of
21 NO3" among streams. There also remains some question of whether NO3" produced from
22 mineralization none-the-less results from atmospheric deposition because mineralization
23 recycles nitrogen from leaf litter. Mineralization during the winter may simply recycle
24 nitrogen from the leaf fall of the previous autumn; some portion of the nitrogen incorporated
25 into leaves in the summer undoubtedly originates as atmospheric deposition. In addition,
26 chronic nitrogen deposition has probably contributed to forest growth in the past (through
27 fertilization), and nitrogen now being mineralized may be the result of such "excess" storage
28 of nitrogen in forest biomass.
29 Earlier in this document (see Section 10.6.2.3) it was suggested that the severity and
30 duration of NO3" episodes can be expected to increase as forests become more nitrogen
31 sufficient (see also Driscoll and Schaefer, 1989; Stoddard and Murdoch, 1991). Some of the
August 1991
10-160 DRAFT-DO NOT QUOTE OR CITE
-------
1 best information on whether atmospheric deposition is contributing to NO3" episodes may
2 therefore come from an examination of long-term trends in surface water NO3"
3 concentrations.
4 There is some evidence that the occurrence and severity of NO3" episodes are
5 increasing. Smith et al. (1987a) examined trends in NO3" at 383 stream locations in the
6 United States between 1974 and 1981, and reported increases at 167 sites, especially east of
7 the 100th meridian. Many of the increasing trends could be attributed to increased use of
8 fertilizers in agricultural areas, particularly in the Midwest. In addition to agricultural
9 runoff, Smith et al. (1987a) identify atmospheric deposition as a major source of NO3" in
10 surface waters, particularly in forested basins of the East (e.g., New England and the Mid-
11 Atlantic) and Upper Midwest. Despite widespread use of fertilizers in most of the regions
12 covered by the Smith et' al. study, they found a high degree of correlation between stream
13 basin yield of NO3" and rates of nitrogen deposition.
14 Historical data are available from 19 large streams in the Catskill Mountains, some of
15 which have been monitored since early in this century (Stoddard and Murdoch, 1991;
16 Stoddard, in review). Trend analyses indicate that NO3" concentrations have increased in all
17 of the streams (Table 10-24), with the majority of the increase occurring in the past two
18 decades (Murdoch and Stoddard, in review; Stoddard, in review). These increases are not
19 attributable to other anthropogenic sources of nitrogen, and are similar to trends observed in
20 8 headwaters streams monitored in the 1980s (Murdoch and Stoddard, in press; Murdoch and
21 Stoddard, in review). At four historical Catskill sites where stream discharge data are
22 available,' the relationships between NO3" concentration and discharge have changed over the
23 course of the past 4 decades (Figure 10-30). In all cases, the relationships are steeper in the
24 1980s than in the past, indicating that most of the increase in NO3" has occurred at high
25 flows (i.e., episodic NO3" concentrations have increased more than baseflow NO3"
26 concentrations).
27 Trends in lake'water NO3" concentrations that are similar to the Catskill stream trends
28 have been reported for Adirondack lakes (Table 10-25; Driscoll et al., in review). Nine out
29 of seventeen Adirondack lakes exhibited significant increases in NO3~ concentrations, while
30 only one exhibited a significant decrease (Table 10-25). It is not statistically possible to
31 determine whether episodic NO3" concentrations are mostly responsible for the trends in
August 1991
10-161
DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-24. SLOPES OF NITRATE TRENDS (/Lteq-JL^-yr^) IN CATSKILL
STREAMS BEFORE 1945, BETWEEN 1945 AND 1970, AND AFTER 1970. SLOPES
FOR EACH PERIOD ARE CALCULATED FROM BEST-FIT REGRESSION LINES
(ANALYSIS OF COVARIANCE ON RANKS, SEE TEXT FOR DETAILS) FITTED TO
DATA FROM THE ENTIRE PERIOD OF RECORD. ALL TRENDS ARE
SIGNIFICANT AT P <0.05. MEDIAN VALUES AND SAMPLE SIZES FOR EACH
PERIOD ARE GIVEN IN PARENTHESES. [-- = DATA INSUFFICIENT FOR
ANALYSIS.] FROM MURDOCH AND STODDARD (IN REVIEW)
Site
Batavia Kill
Bear Kill above Grand Gorgeb
Bear Kill above Hardenbergh Falls
Beaver Kill"
Birch Creek above Pine Hill
Birch Creek at Pine Hill
Bush Kill •_
Eushncllville Creek8
Esopus Creek above Big Indian
Esopus Creek below Big Indian
Esopus Creek at Coldbrook
- ,
Little Beaver Kill8
,
Manor Kill
Nevcrsink River
Rondout Creek
Schohario Creek at Praltsville
Stony Clove Creek"
West Kill
Woodland Creek8
Before 1945
+0.24
(11, n=235)
_
+0.34
(18, n=253)
+0.05
(4, n=270)
—
-0.01
(11, n=287)
+0.11
(4, n=235)
+0.04
(4, n=267)
+0.08
(4, n=246)
-0.16
;(7, n=59)
+0.24
(7, n=352)
+0.00
(4, n=268)
-0.12
(11, n=251)
_
—
+0.64
(7,n=238)
-0.00
(4, n=272)
+0.19
(7, n=227)
+0.02
(4, n=272)
Change in Nitrate Concentration
-'• '1945-1970' , "
' +0.21
_
(27, n=9)
_
+0.10
+0.60
:>, , (4, n=12)
+0.68
(6, n=ll)
' +0.00
•(7, n=248)
+0.25
-0.01
(7, n=64)
-0.08
(11, n=784)
+0.01 • -
'' - -0.55
(14, n=306)
+0.33
(7, n=185)
+0.00
(7,n=12)
-0.13 ,'-.-,
(14, n=712)
+0.08
•
+0.08
After 1970
+0.28
(21, n=70)
+0.70
(38,n=92)
_
+ 1.76
(14, n=10)
+2.68
(16, n=75)
+0.73
(19 n=63)
+2.28
(19, n=94)
+ 1.57
(17, n=10) ,
_
+1:98
(21, n=93)
+2.00
(19, n=886)
+0.85
(5, n=10)
+0.97
(17, n=96)
+ 1.28
(14, n=104)
+ 1.79
(8, n=43)
, + 1,93
(21, n=805)
+3.77 "
(24, n=10)
. +3.95
(25, n=10)
*0aia for these sites are available only for periods before 1945 and from 1977-79. Trends reported for the periods of missing data are
based on regression lines for the entire data set; median values cannot be listed.
Data available, for fewer than 2 years in one or more time periods at this site. Trends were not calculated during these time periods at this
site, but median values and sample sizes are listed. .
August 1991
10-162 DRAFT-DO NOT QUOTE OR CITE
-------
Schoharie Creek at Prattsville
Neversink River at Claryvllle
1 10 100
Stream Discharge (m sec )
100
Stream Discharge (m sec )
Esopus Creek at Coldbrook
11 ml 1—i i 11 nil—^—i i i mil
Rondout Cr«ek at Lowes Cornere
19SO-S9
Stream Discharge
Stream Discharge (rn sec )
Figure 10-30. Relationship between nitrate concentration and stream discharge for four
Catskill streams during four most recent decades: (a) Schoharie Creek at
Prattsville, (b) Neversink River at Prattsville, (c) Rondout Creek at
Lowes Corners, and (d) Esopus Creek at Coldbrook. Regression lines for
each decade are from least-squares regression of concentration on the log
of stream discharge, and all regressions are significant (p < 0.05). All
sites indicate that NO3" concentrations at high discharges are higher in
the 1970s and 1980s than in previous decades. From Murdoch and
Stoddard (in review).
1 Adirondack lakes, because the data record is short (1982-89). Plots of temporal NO3"
2 patterns, however, suggest that baseflow values are relatively unchanged, while spring values
3 are increasing (Figure 10-31).
4 A cautionary note in the interpretation of long-term nitrogen trends is introduced by
5 examination of long-term data from streams at the Hubbard Brook Experimental Forest
6 (HBEF). Data from control Watershed #6 through 1977 suggested a strongly increasing trend
August 1991
10-163
DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-25. TRENDS IN NITRATE CONCENTRATIONS FOR ADIRONDACK
LONG-TERM MONITORING LAKES. SLOPES ARE CALCULATED FROM
BEST-FIT REGRESSION LINES (USING ANCOVA ON RANKS; LOFTIS et aL, 1989)
FITTED TO DATA. DATA ARE FROM DRISCOLL et al. (IN REVIEW)
Lake Name
Arbutus Lake
Barnes Lake
Big Moose Lake
Black Lake
Bubb Lake
Cascade Lake
Clear Pond
Constable Pond
Dart Lake
Heart Lake
Lake Rondaxe
Little Echo Pond
Moss Lake
Otter Pond
Squash Pond
West Pond
Windfall Lake
na
96
51
105
104
88
105
104
106
88
103
88
84
105
93
100
106
88
Change in NO3-b
Ofeq-L^-yr'1)
+ 1.05
+0.03
+0.16
+0.04
-0.11
-0.50
+0.51
+ 1.26
+0.34
+0.88
+0.18 " ;
+0.01
0.00
+ 1.50
' +1.14
+0.09
-0.14
Pc
< 0.0001
0.69
0.36
0.79
0.53
0.04
< 0.0001
0.0003
, 0.07
< 0.0001
0.04
0.12
0.94 -
< 0.0001
0.08
0.56
0.82
"Number of individual observations; the period of record for most sites is from June, 1982, to August, 1989.
bSlope of ANCOVA model. Positive slope indicates an increase in NO3", negative number indicates decrease.
"Significance of regression coefficient for date in ANCOVA model.
August 1991
10-164 DRAFT-DO NOT QUOTE OR CITE
-------
cr
v
a.
n
o
- - TREND IN ALL DMA
TREND IN SPWNG DATA
L
cr
1
- - TREND W ALL DATA
TREND IN SPRING DATA
1982 1983 1984 1985 1986 1987 1988 1989
Figure 10-31. Temporal patterns in lake water NO3" concentration for two Adirondack
lakes, (a) Constable Pond, and (b) Heart Lake. Both sites exhibit
increasing trends in NO3" (Table 10-25). The strongly seasonal behavior
of NO3" hi these lakes suggests that most of the increase has occurred in
spring episodic NO3" concentrations.
1 in NO3" (Schindler, 1987), and have been used to suggest that the HBEF watersheds are
2 undergoing nitrogen saturation (Agren and Bosatta, 1988). Examination of the entire 23-year
3 record (1965-1983) from Watershed #6, however, shows no long-term trend (Likens, 1985;
4 Driscoll et al., 1989a), and emphasizes the importance of examining nitrogen processes in a
August 1991
10-165 DRAFT-DO NOT QUOTE OR CITE
-------
1 truly long-term context. Pools of nitrogen associated with soils and forests at HBEF, and
2 elsewhere, are very large (ca. 340,000 moles/ha at HBEF, up to 520,000 moles/ha at other
3 sites in the eastern United States; Federer et al., 1989) and long-lived (the turnover rate for
4 nitrogen at HBEF is estimated at 80 years); small changes in the long-term cycling of
5 nitrogen within this system will have profound effects on stream water chemistry (Driscoll
6 et al., 1989a). While the data reported here for the Catskills can be considered truly
7 long-term (up to 65 years of record), data for the Adirondacks (Driscoll et al., in review) and
8 other areas of the United States (Smith et al., 1987a) span only one to two decades, and
9 should be interpreted with caution.
10 Many of the data discussed above suggest that NO3" episodes are more severe now than
11 they were in the past. These surface water nitrogen increases have occurred at a time when
12 nitrogen deposition has been relatively unchanged in the northeastern United States (Husar,
13 1986; Simpson and Olsen, 1990; Bowersox et al., 1990). If we accept the idea that an
14 increase in the occurrence of NO3" episodes is evidence that nitrogen saturation of watersheds
15 is progressing, then current data suggest that current levels of nitrogen deposition
16 (350-700 eq • ha"1 • yr"1) are too high the for the long-term health of aquatic systems in the
17 Adirondacks, the Catskills, and possibly elsewhere in the Northeast. It is important to note
18 that this supposition is dependent on our acceptance of NO3" episodes as evidence of nitrogen
19 saturation. At this point no measurements of changes in nitrogen cycling have been made to
20 support this.
21 Similar logic would suggest that levels of nitrogen deposition in the Sierra Nevada
22 (ca. 150 eq • ha"1 • yr"1) may be at the upper limit of the levels that would be protective of
23 the long-term health of sensitive aquatic systems in the West. The discrepancy between the
24 levels of nitrogen deposition that produce signs of nitrogen saturation in the Northeast and the
25 West is a good illustration of the need to set deposition levels in terms of a "critical load" to
26 specific systems. The deposition levels measured in the eastern and western United States are
27 within the range of nitrogen critical loads (210 to 1,000 eq • ha"1 • yr"1) suggested by
28 European work in regions of silicate soils of varying sensitivity (Schulze et al., 1989). The
29 Northeast, because of deeper soils and aggrading forests, may be able to absorb higher rates
30 of deposition without serious damage than areas of the mountainous West, where soils are
31 thin and forests are often absent. The abilities of these regions to absorb nitrogen is a
August 1991
10-166
DRAFT-DO NOT QUOTE OR CITE
-------
1 function of the capacities of their watersheds to retain nitrogen. Because these capacities
2 differ from one region to another, the critical loads of nitrogen that will produce signs of
3 degradation also vary from region to region. These differences are at the heart of the critical
4 loads concept of setting deposition limits.
5 , . ....
6 10.6.3.3 Biological Effects
7 . .. In addition to affecting acid/base chemistry, there is considerable evidence that episodic
8 events affect biological systems. J. Baker et al. (1989) have reviewed the scientific evidence
9 of biological effects attributable to surface water acidification, and have documented evidence
10 of: (1) fish kills during spring episodes in Norwegian rivers; (2) loss of stocked trout from
11 re-apidifying Adirondack lakes after liming; (3) decreased density of acid-sensitive benthic
12 invertebrates in whole stream experiments simulating episodic acidification; and (4) loss of
13 acid-sensitive mayfly and stonefly species from two Ontario stream sites experiencing acidic
14 . • • .episodes.. , . ,
15 Although biological responses to acidification are complex,, research over the past
16 , decade has provided us with a strong understanding of the relationship between changes in
17 surface water chemistry associated with acidification and responses in biological communities.
18 In short, shifts in water chemistry during acidification can affept biological communities and
19 .processes both directly (e.g., physiological stress, toxicity) and indirectly (e.g., changes in
20 food availability, predation). From a biological perspective, changes in several ions (e.g.,
21 hydrogen, aluminum, and calcium), not just pH, are important. The scientific literature
22 contains a number of comprehensive reviews of acidification effects on aquatic biota (cf.,
23 National Research Council Canada, 1981; Altshuller and Linthurst, 1984; Baker and
24 ..;•• Christensen, 1991; J. Baker et al., 1989); interested readers are referred to these documents
25 for detailed description of the state-of-science.
26 ; , . . •
27 10.6.4 The Effects of Nitrogen Deposition on Eutrophication
28 The term "eutrophy" generally refers to a state of nutrient enrichment (Wetzel, 1983),
29 but is commonly used to refer to conditions of increased algal biomass and productivity,
30 presence of nuisance algal populations, and a decrease in oxygen availability for heterotrophic
31 organisms. "Eutrophication" is the process whereby lakes, estuaries and marine systems
August 1991
10-167
DRAFT-DO NOT QUOTE OR CITE
-------
1 progress toward a state of eutrophy. In lakes, eutrophication is often considered to be a
2 natural process, progressing gradually over the long-term evolution of lakes. The process can
3 be significantly accelerated by the additional input of nutrients from anthropogenic sources.
4 The subject of eutrophication has been extensively reviewed by Hutchinson (1973), the
5 National Research Council (1969), and Likens (1972).
6 Establishing a link between nitrogen deposition and the eutrophication of aquatic
7 systems depends on a determination of two key conditions. The first condition is that the
8 productivity of the system is limited by nitrogen availability. Our current concept of nutrient
9 limitation stems from Liebig's Law of the Minimum (von Liebig, 1840), which can be
10 paraphrased to suggest that, at any single point in time, ecosystem productivity will be
11 limited by whatever necessary environmental element is in shortest supply. When that
12 necessary environmental element is nitrogen, then the system can be said to be nitrogen
13 limited. The second condition is that nitrogen deposition be a major source of nitrogen to the
14 system. In many cases, the supply of nitrogen from deposition is minor when compared to
15 other anthropogenic sources, such as pollution from either point or non-point sources.
16
17 10.6.4.1 Freshwater Eutrophication
18 It is generally accepted that the productivity of fresh waters is limited by the availability
19 of phosphorus, rather than the availability of nitrogen (reviewed by Hecky and Kilham,
20 1988). While conditions of nitrogen limitation do occur in freshwater systems (discussed
21 below), they are often either transitory, or the result of high inputs of phosphorus from
22 anthropogenic sources. At high rates of phosphorus input, phosphorus will cease to be in
23 short supply, and whatever nutrient is then least abundant (often nitrogen) will become
24 limiting. While additions of nitrogen from deposition will lead to increased productivity in
25 these situations, the primary dysfunction is an excess supply of phosphorus, and these
26 situations will not be discussed further. Often when nitrogen limitation does occur it is a
27 short-lived phenomenon, because nitrogen-deficient conditions favor the growth of blue-green
28 algae (e.g., Smith, 1982), many of which are capable of nitrogen fixation. Because
29 nitrogen-fixing species are not limited by the availability of fixed nitrogen (e.g., NH4+,
30 NO3~), they may thrive under conditions where other species are nitrogen limited, and
31 effectively increase rates of nitrogen input to the system by fixation of gaseous nitrogen.
August 1991
10-168 DRAFT-DO NOT QUOTE OR CITE
-------
1 High rates of nitrogen fixation may lead to situations where nitrogen can no longer be said to
2 be limiting, and the system often returns to a state of phosphorus limitation; In lakes,
3 nitrogen fixation may be considered a natural mechanism which compensates for deficiencies
4 in nitrogen, and contributes to the long-term evolution and ubiquity of phosphorus limitation
5 (Schindler, 1977).
6 Nitrogen limitation can occur naturally (i.e., in the absence of anthropogenic
7 phosphorus inputs) in lakes with very low concentrations of both nitrogen and phosphorus, as
8 are common in the western United States, and in the Northeast (Suttle and Harrison, 1988).
9 Suttle and Harrison (1988) and Stockner and Shortreed (1988) have suggested that phosphorus
10 concentrations are too low in these systems to allow blue-green algae to thrive, because they
11 are poor competitors for phosphorus at very low concentrations (e.g., Schindler et al., 1980;
12 Smith and Kalff, 1982). Thus, diatom communities dominate phytoplankton and periphyton
13 communities in these extremely nutrient-poor (ultraoligotrophic) systems, and rates of
14 nitrogen fixation do not increase because blue-green algae do not become established,
15 regardless of relative nitrogen or phosphorus deficiency. In these systems, the two nutrients
16 are often closely coupled and constant shifts between nitrogen and phosphorus deficiency may
17 occur without obvious changes in community structure. In these situations, additional loading
18 of nitrogen from anthropogenic deposition is likely to have only a small effect on primary
19 productivity because the system quickly becomes phosphorus limited. In a literature survey
20 of 62 separate nutrient limitation studies in lakes, Elser et al. (1990) found that simultaneous
21 additions of nitrogen and phosphorus produced the largest growth response in 82% of the
22 experiments. These results underline the likelihood that a lake limited by one nutrient may
23 quickly become limited by another if the lake becomes enriched with the original limiting
24 nutrient.
25 Estimations of nutrient limitation in lake ecosystems follow three major lines of
26 reasoning: (1) evidence from ambient nutrient concentrations and the nutritional needs of
27 algae; (2) evidence from bioassay experiments at various scales; and (3) evidence from
28 nutrient dynamics and input/output studies (Hecky and Kilham, 1988; Howarth, 1988).
29 Much of the acceptance of the idea that freshwater lakes are primarily phosphorus
30 limited stems from the close correlations between phosphorus concentrations and lake
31 productivity or algal biomass (usually measured as chlorophyll concentration) that have been
August 1991
10-169 DRAFT-DO NOT QUOTE OR CITE
-------
1 observed in a large number of lake studies (e.g., Dillon and Rigler, 1974; Schindler, 1977,
2 1978; reviewed by Reynolds, 1984; Peters, 1986). More recently, researchers have begun to
3 question the ubiquity of the phosphonisxhlorophyll relationship, and to identify some of the
4 factors that lead to the large variability observed in this relationship in nature (e.g., Smith
5 and Shapiro, 1981; Smith, 1982; Pace, 1984; Hoyer and Jones, 1983; Prairie et al., 1989).
6 Notably, researchers have found that the relationship is not linear, as previously supposed,
7 but sigmoidal (McCauley et al., 1989), and that the slope of the relationship is significantly
8 affected by nitrogen concentrations, particularly at high concentrations of phosphorus
9 (> 10 /ieq • L"1) that are likely to be caused by anthropogenic inputs. McCauley et al.
10 (1989) found that nitrogen had little effect on the phosphorus:chlorophyll relationship at low
11 concentrations of phosphorus. This effect is expected in nutrient-poor lakes where the
12 primary effect of nitrogen additions would be to push lakes into a phosphorus-deficient
13 condition.
14 Arguments based on ambient nutrient concentrations stem from the early work of
15 Redfield (1934), who examined the concentrations of nutrients within the cells of nutrient-
16 sufficient algae from marine systems worldwide, and found surprisingly consistent results for
17 the ratio of carbon to nitrogen to phosphorus concentrations (106:16:1); deviations from these
18 ratios are taken to be evidence that one nutrient or another is limiting to algal growth (e.g.,
19 N:P ratio values below 16:1 suggest nitrogen limitation; values above 16:1 suggest
20 phosphorus limitation). Because the relative supply rates of phosphorus and nitrogen will
21 determine whether one or the other nutrient is in short supply, it has been suggested that the
22 ratio of the two nutrients (i.e., total nitrogen:total phosphorus) can be used as an index of
23 nutrient limitation (Chiandani and Vighi, 1974; Rhee, 1978; Schindler, 1976, 1977, 1978).
24 Various researchers have extended interpretation of the Redfield ratio to include ambient
25 nutrient concentrations in water (Redfield's original work was with intracellular
26 concentrations), and applied the nutrient ratio criteria to waters supplying lakes to determine
27 the likely limiting conditions that these waters will produce (e.g., Schindler, 1977; Smith and
28 Shapiro, 1981; Prairie et al., 1989). This method has the potential to illustrate regional
29 patterns and has gained some support from the results of bioassay experiments (see below).
30 This idea has been refined recently to exclude from the ratio those forms of nitrogen and
31 phosphorus that are not biologically available (e.g., especially organic forms of nitrogen),
August 1991
10-170 DRAFT-DO NOT QUOTE OR CITE
-------
1 with the result that good predictions of nutrient limitation can now be made from ratios of
2 total inorganic nitrogen (DIN) to total phosphorus (TP) (Morris and Lewis, 1988).
3 Morris and Lewis (1988) conducted nutrient addition bioassays on natural assemblages
4 of phytoplankton from many lakes, and compared their results to DIN:TP values measured in
5 the lakes at the same time as the experiments were conducted. They found that lakes with
6 DINrTP values less than 9 (using molar concentrations) could be limited by either nitrogen or
7 phosphorus (often additions of both nutrients were required to stimulate growth), while lakes
8 with DIN:TP values less than 2 were always limited by nitrogen. The discrepancy between
9 the 16:1 Redfield ratio, and the 2:1 ratio suggested by Morris and Lewis (1988), may result
10 from measuring ambient, rather and cellular, nutrient concentrations, and from the variety of
11 critical N:P ratios exhibited by different species in nature (Suttle and Harrison, 1988).
12 If a critical DIN:TP value less than 2 is applied to lakes from the Eastern Lake Survey
13 (Linthurst et al., 1986) and Western Lake Survey (Landers et al., 1987), it is possible to
14 estimate the number of nitrogen limited lakes in some regions of the United States
15 (Table 10-26). Lakes with total phosphorus concentrations greater than 2.0 /zeq • L"1 have
16 been excluded from this analysis because many of them may have experienced anthropogenic
17 inputs of phosphorus (Vollenweider, 1968; Wetzel, 1983). This test is therefore a
18 conservative one for nitrogen limitation, both because the DINrTP value chosen (< 2) is a
19 conservative measure of nitrogen limitation (Morris and Lewis, 1988), and because some
20 lakes with naturally high concentrations of phosphorus may be excluded; these lakes are more
21 likely to be nitrogen-limited than lakes with low phosphorus concentrations. The proportions
22 Of lakes that can be considered nitrogen-limited vary widely from region to region, with the
23 greatest number being found, as expected, in the West. The highest proportion was found in
24 the Pacific Northwest (27.7% of lakes exhibited low DINrTP ratios), but all sub-regions of
25 the West contained substantial numbers of nitrogen-limited lakes. The smallest proportions
26 were found in the Southeast (2.5% of the lakes in the entire region exhibited low DIN:TP
27 ratios) and the Northeast (5%). One surprise in this analysis is the number of lakes in the
28 Upper Midwest that appear to be nitrogen-limited; taken as a whole, this region had 19% of
29 its lakes with DIN: TP ratios less than one.
30 A more direct indication of nutrient limitation than is available from nutrient ratios can
31 be gained from bioassay experiments, where a small volume of natural lake water is
August 1991
10-171
DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-26. ESTIMATED NUMBER AND PROPORTION OF NITROGEN-
LIMITED LAKES IN SUBREGIONS OF THE UNITED STATES SAMPLED BY THE
NATIONAL SURFACE WATER SURVEY. ESTIMATES ARE BASED ON MOLAR
RATIOS OF TOTAL INORGANIC NITROGEN CONCENTRATIONS (NITRATE +
AMMONIUM) TO TOTAL PHOSPHORUS CONCENTRATIONS
Sub-Region
Eastern Lake Surveya
Adirondacks (1A)
Poconos/CatsMUs (IB)
Central New England (1C)
Southern New England (ID)
Northern New England (IE)
Northeastern Minnesota (2A)
Upper Peninsula, Michigan (2B)
Northcentral Wisconsin (2C)
Upper Great Lakes Area (2D)
Southern Blue Ridge (3A)
Florida (3B)
Western Lake Surveyb
California (4A)
Pacific Northwest (4B)
Northern Rockies (4C)
Central Rockies (4D)
Southern Rockies (4E)
Number of
Lakes in
Sub-Region
1,684
1,986
2,003
2,667
2,388
2,132
1,698
1,707
6,147
538
8,053
2,806
2,200
3,335
2,970
2,195
Estimated Number
of Nitrogen
Limited Lakes
16.4
228.5
54.9
144.7
91.3
316.2
305.8
248.2
1345.4
11.5
2.5
535.8
609.1
739.9
788.7
455.2
Proportion
of Population
N-Limited (%)
1.0
11.5
2.7
5.4
3.8
14.8
18.0
14.5
21.9
2.1
0.0
19.1
27.7
22.2
26.6
20.7
"Data from Kanciruk et al. (1986); excluding lakes with total phosphorus > 2 //mol • L'.
"Data from Eilers et al. (1987); excluding lakes with total phosphorus > 2 jumol • L"1.
August 1991
10-172 DRAFT-DO NOT QUOTE OR CITE
-------
1 enclosed, and various known concentrations of potentially limiting nutrients are added (e.g.,
2 Melack etal., 1982; Setaro and Melack, 1984; Stoddard, 1987b). A growth response
3 (usually measured as an increase in biomass) in treatments containing an added nutrient
4 constitutes evidence of limitation by that nutrient. The results of such experiments are
5 available for only a.few selected nutrient-poor lakes, however, and indicate a variety of
6 responses including strong phosphorus limitation (Melack et al., 1987), limitation by
7 phosphorus and iron (Stoddard, 1987b), simultaneous nitrogen and phosphorus limitation
8 (i.e., the two nutrients are so closely balanced that addition of one alone simply leads to
9 limitation by the other, Gerhart and Likens, 1975; Suttle and Harrison, 1988; Dodds and
10 Priscu, 1990), and limitation primarily by nitrogen (Morris and Lewis, 1988; Goldman,
11 1988). No clear pattern of nitrogen or phosphorus limitation develops from an examination
12 of these few studies. .
13 The potential for nitrogen deposition to contribute to the eutrophication of freshwater
14 lakes is probably quite limited. Eutrophication by nitrogen inputs will only be a concern in
15 lakes that are chronically nitrogen-limited. This condition occurs in some lakes that receive
16 substantial inputs of anthropogenic phosphorus, and in many lakes where both phosphorus
17 and nitrogen are found in low concentrations (e.g., Table 10-26). In the former case, the
18 primary dysfunction of the lakes is an excess supply of phosphorus, and controlling nitrogen
19 deposition would be an ineffective method of water quality improvement. In the latter case,
20 the potential for eutrophication by nitrogen addition (e.g., from deposition) is limited by low
21 phosphorus concentrations; additions of nitrogen to these systems would soon lead to
22 nitrogen-sufficient, and phosphorus-deficient, conditions. Increases in nitrogen deposition to
23 some of the regions in Table 10-25 would probably lead to measurable increases in algal
24 biomass in those lakes with low DIN:TP ratios and substantial total phosphorus
25 concentrations, but the number of lakes that meet these criteria is likely to be quite small.
26
27 10.6.4.2 Estuaries and Coastal Waters
28 Estuarine and coastal water ecosystems exist at the transition between freshwater
29 systems and the open ocean. These transition zones share some characteristics with
30 freshwater and marine systems, but they also have some unique properties that lead to
31 different responses to nitrogen oxide deposition and a correspondingly different set of
August 1991
10-173 DRAFT-DO NOT QUOTE OR CITE
-------
1 concerns. They are at the end of a long series of nitrogen transport and transformation
2 processes involving interactions with vegetation, soils, groundwater, small streams, lakes, and
3 rivers. At each step in this series the processes vary temporally and spatially and may be
4 subject to a variety of human influences. This transition zone integrates complex and
5 fluctuating processes which are distributed over what are sometimes very large watersheds.
6 The transition zones between fresh and salt water systems are subject to natural
7 processes which are not observed elsewhere in aquatic systems, such as tidal flows and
8 salinity changes. They are also subject to substantial human influence. Estuaries provided
9 natural ports and are among the most productive ecosystems on the planet (Begon et al.,
10 1986). Thus they became an obvious location for cities, with accompanying demands for
11 waste water disposal. The history of human use of estuaries and lands around estuaries make
12 it more difficult to isolate the effects of a particular anthropogenic contaminant on ecosystem
13 characteristics. The conservative approach used above to assess the impact of nitrogen
14 deposition on freshwater eutrophication (excluding all systems with anthropogenic impacts
15 other than atmospheric deposition) is not possible for estuaries and coastal waters; all
16 estuarine systems, and most coastal waters, have been subjected to human impacts, often for
17 several centuries.
18 Estuaries are bodies of water, more or less isolated from the rest of the ocean, where
19 fresh water and salt water mix. This generally produces a salinity gradient* and often leads
20 to stratification of water with the heavier salt water below a layer of fresh water. Estuaries
21 are also subject to tidal effects and may be strongly influenced by river flows. In
22 combination these forces tend to produce quite complex water circulation patterns with
23 significant biological consequences. For example, water currents within Chesapeake Bay
24 concentrate and circulate the dinoflagellate, Gyrodinium uncatenum, responsible for red tides
25 in that estuary (Tyler etal., 1982). Circulation patterns within estuaries may also influence
26 patterns of habitat use by fish (e.g., Pietrafesa et al., 1986).
27 Boynton et al. (1982) described a classification of estuaries into four categories that
28 were designed to reflect the primary factors influencing algal production, and the variability
29 which exists among estuaries:
30 • Fjords have deep basin waters and shallow underwater sills connecting them with the
31 sea, providing slow exchange with adjacent sea waters;
32
August 1991
5 10-174 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44
• Lagoons are shallow, well-mixed, slowly flushed, and only slightly influenced by
riverine inputs;
• Embayments are deeper than lagoons, often stratified, only slightly influenced by
freshwater inputs, and have good exchange with the ocean; and
• River-Dominated Estuaries are a more diverse group of systems, all of which exhibit
seasonally depressed salinities due to riverine inputs and variable degrees of
stratification! '
The physical and chemical structure of estuaries will strongly shape the movement and
transformation of nitrogen compounds» Aston (1980) has provided a,list of features of
. estuaries which have a controlling influence on the geochemistry of contaminants and
nutrients: , ,: . : •-...• ,;
(a) The tidal mixing of fresh and sea waters on a semi-diurnal or diurnal time scale, with
corresponding changes in the volume of water in an estuary, produces temporal changes
in the contributions of nutrients 'and dissolved gases from marine and fresh water
sources. For example, estuaries are generally enriched in nutrients relative to ocean
waters due to the local influences of land drainage and often pollution.
(b) The circulation, and especially the stratification, of some estuaries can create vertical
and horizontal variations of the concentrations of nutrients and dissolved gases within an
estuary.
(c) Estuarine topography may give rise to particularly restricted circulations (e.g., in fjords,
where the mixing of external sea water with the estuarine waters is greatly reduced) and
the restricted mixing leads to unusual chemical environments, for example
oxygen-deficient waters. ,
(d) The circulation patterns in coastal waters and estuaries lead to the deposition of various .
types of sedimentary material. The deposition and resuspension of sediments may
influence the budgets of dissolved constituents, including nutrients' and gases, in
estuarine waters.
(e) Chemical reactions occurring during the mixing of river water with sea water may lead
to the removal or addition of the dissolved nutrients. Also, the changes in temperature
and salinity during estuarine mixing influence the solubility of dissolved gases and thus
influence their removal or addition in an estuary.
(f) Biological production and metabolism have significant influences on the occurrence and
distribution of nutrients and some gases (e.g., carbon dioxide and oxygen) in estuarine
waters. The biological communities in estuaries tend to be species-poor, because few
species are able to tolerate the extremes in environment to, which they .are exposed.
What species do thrive, however, are often productive.
August 1991
10-175 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
S
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
In fact, estuaries may be extremely productive. Fisheries yields in estuaries are higher
per unit area than in lakes (Nixon, 1988). This appears not to be related to primary
production, but rather to the efficiency of utilization of the primary production. The input of
nutrients from outside the ecosystem may be a major determinant of overall fisheries
production levels (Day et al., 1982). The economic importance of estuaries may be simply
indicated by McHugh's (1976) estimate that in 1970, 69% (by weight) of fish landings in the
United States were estuary-dependent.
Estuaries and coastal waters receive substantial amounts of weathered material (and
anthropogenic inputs) from terrestrial ecosystems and from exchange with sea water. As a
' ;'.',•• . ^;
result, they tend to be very well buffered; acidification is not a concern in any of these areas.
The same load of weathered material and anthropogenic inputs that makes estuaries and
coastal areas insensitive to acidification, however, makes them very prone to the effects of
eutrophication. Eutrophication of these areas has some very specific and damaging
consequences) especially the creation of anoxic bottom waters, blooms of nuisance algae, and
replacement of economically-important species by less-desirable ones (e.g., Mearns et al.,
1982; Jaworski, 1981). Eutrophication, for example, has been suggested as the causal factor
i
in the disappearance of the striped bass (Morone saxatilis) fishery in Chesapeake Bay (Price
et al., 1985); the increasing spatial extent of anoxic bottom waters during the summer season
is the proposed mechanism (e.g., Officer et al., 1984). Anoxia is also thought to have had
disastrous effects on surf clams (Spisula solidissima) in the New York Bight (Swanson and
Parker, 1988) and the blue crab (Callinectes sapidus) habitat in Chesapeake Bay (Officer
et al., 1984). In 1971 blooms of the red tide dinoflagellate Ptychodiscus brevis in the Gulf of
Mexico were responsible for the deaths of approximately 100 tons of fish daily; the high
nutrient concentrations typical of eutrophic conditions have been linked to many blooms of
nuisance algae (Paerl, 1988).
Establishing a link between nitrogen deposition and the eutrophication of estuaries and
coastal waters depends on a determination (as it does in freshwater—see above) of two key
conditions. The first condition is that the productivity of these systems is limited by nitrogen
availability. The second condition is that nitrogen deposition be a major source of nitrogen to
the system. In many cases, the supply of nitrogen from deposition is minor when compared
to other anthropogenic sources, such as pollution from either point or non-point sources.
August 1991
10-176 DRAFT-DO NOT QUOTE OR CITE
-------
1 Few topics in Aquatic biology have received as much attention in the past decade as the
2 debate over whether estuarine and coastal ecosystems are limited by nitrogen, phosphorus, or
3 some other factor (reviewed by Hecky and Kilham, 1988). In a seminal paper published in
4 1971, Ryther and Dunstan (1971) used evidence of ambient nutrient concentrations, and the
5 results of bioassay experiments, to conclude that nitrogen limited the productivity of waters
6 along the south shore of Long Island and in the New York Bight. They noted that, during
7 blooms of algae in these areas, inorganic nitrogen concentrations often decreased to levels
8 below detection, while inorganic concentrations of phosphorus remained high. From this
9 evidence they deduced that phosphorus could not be a limiting factor, but that nitrogen could
10 be. They conducted bioassay experiments, suspending in small bottles single-species cultures
11 of either Nannochloris atomus or Skelatonema costatum, the two algal species that were
12 dominant in the blooms in each location, in filtered sea water with additions of either
13 ammonium or phosphorus. Ryther and Dunstan (1971) found that both species increased
14 dramatically in ammonium-enriched bottles, but that phosphorus-enriched bottles were no
15 different than controls, and that this response was consistent at a large number of sites
16 throughout the south shore of Long Island and in the New York Bight. They concluded that
17 "nitrogen is the critical limiting factor to algal growth and eutrophication in coastal marine
18 waters" (Ryther and Dustan, 1971, p. 1008).
19 Since the publication of this influential paper, many researchers have accepted the
20 notion that coastal waters and estuaries are limited primarily by nitrogen (e.g., Boynton
21 et al., 1982; Nixon and Pilson, 1983), to the point where nitrogen-limitation in marine
22 waters, and phosphorus-limita.tion in freshwaters, has become near dogma (Hecky and
23 Kilham, 1988). More recently, some oceanographers have begun to question the ubiquity of
24 nitrogen-limitation in estuarine and coastal marine waters (e.g., Smith, 1984; Howarth,
25 1988), and it seems clear that evidence for nutrient limitation in these systems must be
26 analyzed on a case-by-case basis. Experiments to confirm widespread nitrogen limitation in
27 estuaries have not been conducted, and nitrogen limitation can not be assumed to be the rule
28 (Hecky and Kilham, 1988).
29 Estimations of nutrient limitation in estuaries and coastal marine ecosystems follow the
30 same three major lines of reasoning as arguments about freshwater nutrient limitation (see
31 Section 10.6.4.1): (1) evidence from ambient nutrient concentrations and the nutritional
August 1991
10-177 DRAFT-DO NOT QUOTE OR CITE
-------
1 needs of algae; (2) evidence from bioassay experiments at various scales; and (3) evidence
' • , ' ' p • • f
2 from nutrient dynamics and input/output studies (Hecky and Kilham, 1988; Howarth, 1988).
3 As explained earlier, arguments based on ambient nutrient concentrations stem from the
4 early work of Redfield (1934), who examined the concentrations of nutrients within the cells
5 of nutrient-sufficient algae from marine systems worldwide, and found surprisingly consistent
6 results for the ratio of carbon to nitrogen to phosphorus concentrations (106:16:1, using
7 molar concentrations); deviations from these ratios are taken to be evidence that one nutrient
8 or another is limiting to algal growth (e.g., molar N:P ratio values below 16:1 suggest
9 nitrogen-limitation; values above 16:1 suggest phosphorus-limitation). Various researchers
10 have extended interpretation of the Redfield ratio to include ambient nutrient concentrations in
11 water (Redfield's original work was with intracellular concentrations), and applied the
12 nutrient ratio criteria to waters supplying estuaries and coastal systems to determine the likely
13 limiting conditions that these waters will produce (e.g., Ryther and Dunstan, 1971; Jaworski,
14 1981). The biotic response (i.e., biostimulation) is not measured using this approach, but is
15 instead inferred from geochemical principles; in this sense, the nutrient ratio approach
16 measures potential nutrient limitation rather than actual limitation. Boynton et al. (1982)
17 summarized nutrient ratio information for a number of estuarine systems; these results are
18 repeated in Table 10-27. At the time of maximum primary productivity, a majority of the
19 estuaries they surveyed (22 out of 27) had N:P ratios well below the Redfield ratio and may
20 have been nitrogen-limited.
21 The data in Table 10-27, as well as from many other studies, suggest that N:P ratios
22 vary widely within a single system from season to season. D'Elia et al. (1986), for example,
23 report ratios for the Patuxent River estuary that vary from over 20:1 during the winter, to
24 less than 1:1 during the summer. This variability suggests that estuarine algae may be
25 limited by different nutrients at different seasons.
26 The ambient nutrient ratio approach has been criticized widely because it ignores several
27 factors known to be important to algal growth. The use of only inorganic nutrient species in
28 the ratios, for example, has been criticized because many algal species are known to utilize
29 organic forms, especially of phosphorus (Howarth, 1988); the nutrient ratios listed for
30 freshwaters systems (see freshwater eutrophication section, above) were based on
31 concentrations of total inorganic nitrogen and total phosphorus because these are thought to
August 1991
10-178 DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-27. MOLAR RATIOS OF DISSOLVED INORGANIC NITROGEN (DIN)
TO DISSOLVED INORGANIC PHOSPHORUS (DIP) IN A VARIETY OF
ESTUARIES. FROM BOYNTON et al. (1982)
Estuary
Pamlico River, NC
Roskeeda Bay, Ireland
Narragansett Bay, RI
Bedford Basin, Nova Scotia
Beaufort Sound, NC
Chincoteague Bay, MD . • •
Western Wadden Sea, Netherlands
Eastern Wadden Sea, Netherlands
Peconic Bay, NY
Mid-Patuxent River, MD
S.E. Kaneohe Bay, HI
St. Margarets Bay, Nova Scotia
Central Kaneohe Bay, HI
Long Island Sound, NY
Lower San Francisco Bay, CA
Upper San Francisco Bay, CA
Barataria Bay, LA
Victoria Harbor, B.C.
Mid-Chesapeake Bay, MD
Duwamish River, WA
Upper Patuxent River, MD
Baltic Sea
Loch Etive, Scotland
Hudson River, NY
Vostock Bay, USSR
Apalachicola Bay, FL
High Venice Lagoon, Italy
DIN:DIP Ratio at
Time of Maximum
Productivity
0.2
' 0.3
0.5
0.8
1.0
. , . .1-2 - .
1.3
1.5
1.5
1.8
2^0
0.2
2.8
3.9
6.0
6.0
6.2
6.2
7.6
8.5
9.2
15
Redfield Ratio N:P = 16:1
18
20
20
20
48 • ' . .
Annual Range
in DIN:DIP Ratio
0-3
0-1
0.5-14
0.5-8
0.5-16
1-10
1.3-120
1.5-56
1-4
1.8-53
Not reported
1-7
Not reported
1-6
4.5-8.5
0.5-16
6-16
6-15
7-225
8-16
9-61
Not reported
12-125
16-30
5-22
5-22
48-190
August 1991
10-179 DRAFT-DO NOT QUOTE OR CITE
-------
1 be better estimators of the nutrient species actually available to algae (Morris and Lewis,
2 1988). Algal growth may also be more dependent on the supply rates of nutrients than on
3 their ambient concentrations (Goldman and Gilbert, 1982; Healey, 1973); many species of
4 algae may therefore not be limited by nutrients whose ambient concentrations are so low as to
5 be undetectable. Broecker and Peng (1982) have echoed the earlier conclusions of Redfield
6 himself (1958) in pointing out that biologically-mediated nitrogen fixation, and loss rates of
7 nitrogen from the surface waters of marine ecosystems, interact with terrestrial nutrient inputs
8 and tend to push the N:P ratio in the particulate (i.e., living) fraction of water toward a
9 "geochemicaUy balanced" ratio (i.e., the Redfield ratio of 16:1). Thus ratios within the
10 biologically-active portion of the ecosystem (particularly the algae) may approach 16:1...
11 despite much lower ratios in the abiotic portion of the ecosystem. Taken as a whole, the
12 evidence for nitrogen limitation from ambient nutrient concentrations in estuaries and coastal
13 waters must be considered equivocal.
14 A second, and more direct, line of evidence for nutrient limitation in estuaries and
15 coastal waters comes from bioassay experiments. These experiments have been conducted in
16 both freshwater and marine systems at a number of scales from small single-species cultures
17 (Level I experiments), to small enclosures of natural algal assemblages (Level II), to
18 intermediate-sized enclosures (mesocosms) of natural assemblages Level III),, to whole-system
19 (so far largely limited to whole lakes) treatments (Level IV; levels as defined by Hecky and
20 Kilham, 1988). These experiments therefore progress along a gradient of "naturalness" from
21 studies substantially different from the real world (Level I), to those that simulate natural
22 conditions very closely (Levels III and IV). Interpretation of the results of these experiments
23 therefore follows the same gradient, with more confidence being placed in the results of
24 studies at the upper (i.e., more natural) end of the gradient (Hecky and Kilham, 1988;
25 Howarth, 1988). The results of Level I experiments on single-species cultures of algae, like
26 the original experiments of Ryther and Dunstan (1971), are especially difficult to interpret
27 because threshold N:P ratios for individual species are known to vary substantially. Suttle
28 and Harrison (1988) report limitation at ratios from 7:1 to 45:1 for single species. At all
29 scales, the experimental procedure used for experimental nutrient additions is fairly similar,
30 with various nutrients being added either alone or in combination, and the growth in treated
31 enclosures being compared to growth in control enclosures.
August 1991
10-180
DRAFT-DO NOT QUOTE OR CITE
-------
1 Level I and Level II experiments have been conducted in a wide variety of estuaries and
2 coastal waters (e.g., Thomas, 1970; Ryther and Dunstan, 1971; Vince and Valiela, 1973;
3 Smayda, 1974; Goldman, 1976; Graneli, 1978) and often suggest nitrogen limitation. Two
4 studies have suggested seasonal changes from nitrogen limitation to phosphorus limitation
5 (D'Elia et al., 1986; McComb et al., 1981); in both cases nitrogen deficient conditions were
6 found during the peak of annual productivity in the summer. The results of experiments at
7 Levels I and II suggest that nitrogen limitation is at least a common, if not ubiquitous,
8 phenomenon in coastal and estuarine waters. This interpretation has been challenged by
9 Smith (1984) and Hecky and Kilham (1988) because the experiments were conducted at such
10 an unrealistic spatial scale. In particular, Level I and II experiments measure only the
11 short-term response of algae present at the time the experiments are run; they do not allow
12 natural mechanisms such as species replacement and nitrogen fixation to take place.
13 Only a few examples of Level in bioassays exist for estuarine and coastal ecosystems.
14 The best known of these have been conducted at the Marine Ecosystem Research Laboratory
15 (MERL) at the University of Rhode Island. The MERL tanks are large (13 m3), relatively
16 deep (5 m) cylinders, with natural sediments and filtered seawater inputs. They are designed
17 to mimic the environment of the Narragansett Bay, including the mixing, flushing,
18 temperature, and light regimes (Nixon et al., 1984). In the original experiments conducted in
19 the MERL tanks, nutrients were added with ratios that matched those of sewage entering
20 Narragansett Bay, but at concentrations that ranged from IX to 32X those in the bay itself;
21 the experiments were run for 28 mo. Algal abundance, primarily diatoms, increased with the
22 level of nutrient enrichment, but not on a 1:1 basis. Productivity increased only by a factor
23 of 3.5 in the 32X treatment, suggesting that something other than nutrients was limiting for at
24 least a portion of the experiment (Oviatt et al., 1986). Oviatt et al. (1989) have suggested
25 that, in treatments with high levels of nutrient enrichment, grazing by zooplankton controlled
26 algal abundances to low levels, and that the upper limit to productivity was set by
27 self-shading in the algal community. Further experiments conducted with varying nutrient
28 ratios suggested that diatoms in the low-nutrient (IX) treatments were limited by silica, and
29 not by either nitrogen or phosphorus (Doering et al., 1989). Sewage inputs to many
30 estuaries, including Narragansett Bay, are deficient in silica (Officer and Ryther, 1980), and
31 silica concentrations often fall to very low levels during winter diatom blooms in this area
August 1991
10-181 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
S
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
(Pratt, 1965). Taken as a whole, the results of the MERL experiments suggest a complex
picture for Narragansett Bay, where no nutrient is strongly limiting to algal biomass through
much of the year, and where algal abundances during winter blooms are controlled ultimately
by the concentrations of silica.
In another Level HI bioassay experiment, D'Elia et al. (1986) simulated the
environment of the Patuxent River estuary, a tributary to Chesapeake Bay, in 0.5 m3
enclosures. Their results had a strong seasonal component. Supplements of nitrogen, either
as NO3" or as NH4+, stimulated growth during the low-flow, late-summer season. This
corresponds to the time period when N:P ratios in the estuary are low (1:1 or lower).
Phosphorus additions stimulated growth during the late-winter, high-flow season, when N:P
ratios typically exceed 20:1. Peaks in algal abundance occurred in the summer, when anoxic
conditions in bottom waters in Chesapeake Bay are common, and when algae appear to be
nitrogen-deficient.
Thus far only one Level IV experiment has been conducted in estuarine waters, and
only preliminary results are available. Sewage treatments supplying nutrients to the
Himmerfjard basin, a brackish fjord in the Stockholm archipelago on the eastern coast of
Sweden, have been deliberately altered to produce varying levels of phosphorus and nitrogen
loads since 1983 (Graneli et al., 1990). Between 1983 and 1985, phosphorus removal at the
plant was deliberately reduced to produce a 10-fold increase in orthophosphate, and additional
sewage inputs were routed into the basin to increase total nitrogen inputs by 30-40%. At the
same time as nutrient manipulations were being carried out, measurements were made of
nitrogen cycling in the basin, and algal bioassays were conducted to determine nutrient
limitation. Preliminary results suggest that nitrogen is limiting at low nutrient concentrations
(i.e., typical of near-coastal regions unaffected by anthropogenic inputs), and that limiting
nutrients in areas affected by anthropogenic inputs are determined by the supply ratios of
nitrogen and phosphorus (Graneli et al., 1990). Because small changes in the supply of
either phosphorus or nitrogen in the Himmerfjard basin have caused changes in the identity of
the limiting nutrient (i.e., increases in phosphorus quickly lead to nitrogen limitation, and
vice versa), the authors suggest that management of both nitrogen and phosphorus is
necessary to reduce eutrophication in the basin.
August 1991
10-182 DRAFT-DO NOT QUOTE OR CITE
-------
1 The remaining line of evidence used to infer nutrient limitation in estuarine and coastal
2 marine ecosystems comes from studies of nutrient dynamics, and especially of input/output
3 budgets. In many ways, the results of these studies help to integrate the sometimes
4 contradictory results gleaned from studies of nutrient ratios and bioassay experiments at
5 different levels of complexity. Smith (1984) summarized the studies conducted on four
6 sub-tropical bays and concluded that phosphorus is more likely to be limiting in these systems
7 than nitrogen, and that physical factors are often more important than either nutrient. Smith
8 noted that in the systems that had high through-puts of water (i.e., "embayments" according
9 to Boynton et al.'s [1982] criteria, see earlier description), incoming ratios of nutrients were
10 matched very closely by the ratios in the outgoing water. This suggests that algal growth is
11 having little effect on nutrient levels, and that nutrients do not limit productivity. In systems
12 that flush more slowly (i.e., "lagoons" or "fjords" in Boynton et al.'s [1982] classification),
13 any deficiencies in nitrogen in the incoming water can be made up by nitrogen fixation on the
14 ocean bottom, and phosphorus is therefore more likely to be limiting.
15 The question of why nitrogen deficiencies in marine systems are hot simply made up by
16 nitrogen fixation, as suggested by Smith (1984), is central to the issue of whether estuaries
17 and coastal waters are primarily limited by nitrogen or not. In lakes (see description in
18 freshwater eutrophication section), conditions of nitrogen deficiency often produce blooms of
19 planktonic blue-green algae, which fix atmospheric nitrogen and act to return the algal
20 community to a condition of nitrogen sufficiency (Schindler, 1977; Flett et al., 1980). Only
21 when N:P ratios are extremely low, and blue-green algae are unable to fix enough nitrogen to
22 bring the ratio up to the Redfield proportions do lakes remain nitrogen limited (Howarth
23 et al., 1988a). Why then doesn't the same phenomenon' (nitrogen fixation by blue-greens)
24 occur in nitrogen deficient marine systems? A major difference in the biogeochemistry of
25 lakes and estuaries is that nitrogen fixation by free-living algae (phytoplankton) rarely occurs
26 in estuaries, even when the N:P ratios of incoming water suggest severe nitrogen limitation.
27 Howarth et al. (1988b), for example, surveyed a large number of estuaries along the Atlantic
28 coast of the United States, and found no instances in which nitrogen-fixing blue-green algae
29 made up more than 1 % of the algal biomass. A number of explanations for this lack of
30 nitrogen fixation in estuaries have been proposed, including shorter water residence times
31 (faster flushing rates) than lakes, greater turbulence than in lakes, and lower concentrations of
August 1991
10-183 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
micronutrients (especially iron and molybdenum) needed for the biochemically pathways in
nitrogen fixation (Howarth, 1988; Howarth et al., 1988b). Of these, only the last argument
really holds true in a comparison of lakes and estuaries. Howarth and Cole (1985) and Cole
et al. (1986) have determined that the high concentrations of sulfate in marine systems
interfere with the assimilation of molybdenum by marine algae, and propose that low rates of
molybdenum availability are in turn limiting to rates of nitrogen fixation in many systems.
Molybdenum limitation, however, has not been experimentally demonstrated in many marine
environments. In fact, many nutrient addition bioassays conducted in benthic environments
have shown that the availability of organic matter and of oxygen-depleted microenvironments
tightly control marine microbial nitrogen fixation potentials (Paerl et al., 1987; Paerl and
Prufert, 1987). Because the enzymes needed for nitrogen fixation are readily inactivated by
oxygen, rates of fixation may be limited by energy availability (i.e., the supply of carbon
reductant) and ambient oxygenation. By and large, nitrogen deficient marine waters are
depleted in readily oxidizable organic matter and are well oxygenated. When high rates of
nitrogen fixation do occur in marine systems, they are usually associated with
bottom-dwelling (benthic) algae (Howarth, 1988); these habitats are relatively enriched with
organic matter and support localized oxygen-depleted microenvironments (Paerl et al., 1987).
Iron is also required for nitrogen fixation, and may limit rates of nitrogen fixation in some
freshwater lakes (Wurtsbaugh and Home, 1983); concentrations of iron in seawater are often
much lower than in freshwater, and while little direct evidence of limitation of nitrogen
fixation by low iron concentrations exists, it is certainly a likely condition (Howarth et al.,
1988b). It is difficult at this point in the debate over marine nitrogen fixation to state
anything definitively beyond the fact that nitrogen fixation is not common in marine waters
(Carpenter and Capone, 1983; Howarth et al., 1988a). One possible conclusion from the
debate among researchers in this field (e.g., Howarth et al., 1988b; Paerl et al., 1987) is that
planktonic nitrogen fixers may be limited by micronutrient availability, while benthic nitrogen
fixers are limited by availability of organic carbon and high ambient oxygen levels, but both
factors, as well as others, probably operate in both environments. Light, for example,
appears to play a role in clear, tropical lagoons (Potts and Whitton, 1977; Wiebe et al.,
1975), because benthic nitrogen-fixing algae in these environments require light for
photosynthesis. The presence of benthic nitrogen fixation in Smith's (1984) sub-tropical
August 1991
10-184 DRAFT-DO NOT QUOTE OR CITE
-------
1 lagoons may help explain the apparent contradiction between his predictions of phosphorus
2 limitation, and experimental results suggesting nitrogen limitation, in slowly-flushed systems.
3 Nixon and Pilson (1983) have summarized the results of numerous input/output studies
4 in estuaries and coastal waters, and related the inputs of various nutrients to algal biomass.
5 Their results for nitrogen are repeated in Figure 10-32, and are supported by a similar
6 analysis conducted by Boynton et al. (1982) for algal productivity. The relationship between
7 nitrogen inputs and mean algal biomass in marine systems is certainly much weaker than the
8 relationship between phosphorus and biomass in lakes (e.g., Schindler, 1978), but is
9 none-the-less suggestive of a general pattern of nitrogen-limitation in these systems
10 (Figure 10-32a). Seasonal effects on nutrient ratios, grazing by zooplankton, and physical
11 factors such as light, circulation patterns and turbidity, all lend uncertainty to the relationship.
12 Perhaps the most important aspect of the relationship is the apparent strong dependence of
13 annual maximum chlorophyll concentrations (Figure 10-32b) on nitrogen inputs (r2 = 0.57,
14 p <0.0001). Many of the most severe impacts of eutrophication are experienced during
15 summer algal blooms; these seem to be more strongly dependent on nitrogen than biomass in
16 other seasons (e.g., D'Elia et al., 1986).
17 In summary, there does seem to be confirmatory evidence of nitrogen limitation in
18 many estuarine and coastal marine ecosystems. This conclusion is a general rule, rather than
19 an absolute one, and other limiting factors certainly occur in some locations, and during some
20 seasons. In general, ratios of nitrogen to phosphorus in inputs to estuaries and coastal waters
21 are much lower than in lakes (Hecky and Kilham, 1988; Howarth, 1988), and this probably
22 contributes strongly to the apparent difference between lakes and marine systems in their
23 nutrient limitation. These low ratios, however, result largely from sewage inputs (Ryther and
24 Dunstan, 1971; Jaworski, 1981; Howarth, 1988), and whether atmospheric deposition of
25 nitrogen contributes to eutrophication in these systems will depend strongly on the relative
26 inputs of nitrogen from these two sources. As stated in the introduction to this section, any
27 ' question of negative impacts on estuaries and coastal waters from nilrogen deposition depends
28 both on a determination of nitrogen limitation, and on a determination that atmospheric
29 deposition is a major contributor of nitrogen to these ecosystems.
30 Anthropogenic sources of nitrogen to estuaries and coastal waters include point sources
31 (such as sewage plant outfalls), fertilizer and animal wastes in runoff, and atmospheric
August 1991
10-185 DRAFT-DO NOT QUOTE OR CITE
-------
50-
= 40-
c.
o
x
o
C
C
£
"E
-H-H-
(o) Summer Means
'6*
30
20
10-
10
6«
4*
0.
3*
100
-*-H
100D
0>
x
c
o
175-
(b) Annual Maxima
15C
125-1-
§• 100-.
S 75
50 J-
o.
£ 25--
Q •
10
E*
D* I*
12
15.
14*
100 1000
Nitrogen Input (/u.oiol L~' yr~1)
Figure 10-32. Concentrations of (a) mean algal chlorophyll, and (b) annual maximum
chlorophyll, in the mid region of various estuaries (1-15) and in the
MERL experimental ecosystems (A-G) as a function of the input of
dissolved inorganic nitrogen. 1 - Providence River estuary, Rhode Island;
2 - Narragansett Bay, Rhode Island; 3 - Long Island Sound; 4 - lower
New York Bay; 5 - Delaware Bay; 6 - Patuxent River estuary, Maryland;
7 - Potomac River estuary, Maryland; 8 - Chesapeake Bay; 9 Pamlico
River estuary, North Carolina; 10 - Apalachicola Bay, Florida; 11 -
Mobile Bay, Alabama; 12 - Barataria Bay, Louisiana; 13 - N. San
Francisco Bay, California; 15 - Kaneohe Bay, Hawaii. Doted lines are
from least squares regression, and are included to show strength of
relationships; they do not imply causation. Note change in scale on
vertical axis. Figure redrawn from Nixon and Pilson (1983).
August 1991
10-186 DRAFT-DO NOT QUOTE OR CITE
-------
1 deposition (predominantly due to nitrogen oxides from combustion and ammonium from
2 agricultural activity). Atmospheric deposition may be supplied directly to the surfaces of
3 estuaries or coastal waters, or supplied indirectly to the watershed and subsequently
4 transported to the coast by river flow. As discussed earlier, nitrogen can be deposited in a
5 variety of forms; two of the contentious issues in determining the impact of nitrogen oxides
6 on estuarine ecosystems are estimating the total deposition, and the uncertainty in the relative
7 proportion contributed by the different forms, especially between dry and wet deposition
8 (e.g., Fisher etal., 1988a).
9 Runoff inputs to estuaries may be the most variable of the nitrogen inputs. They vary
10 with watershed area, precipitation rates, land use patterns (especially the use of fertilizer),
11 and rates of atmospheric deposition. Spring runoff represents a major input of nutrients to
12 estuarine and coastal systems. Runoff inputs vary seasonally (e.g., JaworsM, 1981) and from
13 year to year (e.g., Boynton et al., 1982; JaworsM, 1981). Nitrate inputs to estuaries increase
14 markedly during flooding conditions (Biggs and Cronin, 1981), and are at least partially
15 responsible for the finding that nitrogen is less likely to be limiting in the winter and spring
16 than in the summer (above).
17 Point sources of nutrients may be particularly important near urbanized areas. Sewage
18 inputs contribute more than half of the inorganic nitrogen content to a number of major
19 estuaries in the United States: Long Island Sound (67%), New York Bay (82%), Raritan Bay
20 (86%), San Francisco.Bay (73%), and Delaware Bay (50%) (Nixon and Pilson, 1983).
21 Natural and anthropogenic sources of nitrogen to coastal waters may result in the same
22 form of nitrogen (e.g., NO3") being transported by the same route (e.g., river input). Their
23 • effects will therefore be indistinguishable, and it becomes impossible to assign
24 "responsibility" for a problem to a particular source. This has obvious consequences for
25 policy decisions, since, for example, there are many possible regulatory actions which could
26 all result in the reduction of nitrate input to a particular estuary. It may be more cost
27 effective, for example, to increase the efficiency of nitrogen removal in sewage treatment,
28 than to reduce NOX emissions, even if NO3" inputs from atmospheric deposition are
29 increasing.
30 The first published attempt to determine the relative importances of nitrogen from
31 deposition, and nitrogen from runoff, was that of Correll and Ford (1982) for the Rhode
August 1991
10-187 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
River estuary, a tributary to the Chesapeake Bay. Correll and Ford assumed in their analysis
that all atmospheric nitrogen deposited on the watershed was retained, and that the only
atmospheric inputs of nitrogen to the estuary were those that fell directly on the water
surface. This estimate should therefore be considered a lower limit to the importance of
atmospheric deposition, since some terrestrial watersheds do show retention capacities lower
than 100% (see discussion of nitrogen saturation, above). Correll and Ford (1982) conclude
that, on an annual basis, atmospheric and watershed sources of nitrogen to the Rhode River
are approximately equal. During the summer and fall, a period when the Chesapeake Bay
undergoes substantial anoxia, precipitation inputs of nitrogen may slightly exceed those from
watershed runoff. It is important to note that the watershed of the Rhode River estuary is
small relative to the estuary itself (the watershed is less than six times the size of the estuary).
These results should be extrapolated with caution to situations where watershed sizes may be
orders of magnitude larger than those of the waters that drain them. The entire Chesapeake
Bay, for example, is approximately one-fifteenth the size of its watershed, and the relative
importance of nitrogen falling directly on the water surface would therefore be smaller
relative to terrestrial inputs.
Paerl (1985) has determined that NO3" enriched rain, falling on the waters of Bogue
Sound (an embayment), the Continental Slope, and the Gulf Stream (all off the east coast of
North Carolina), increased algal biomass as much as four-fold, and that rain falling directly
on the ocean surface accounted for as much as 10-20% of the volume of water supplied to
these near-coastal areas. More recent work (Paerl et al., 1990) indicates that rainfall
additions as low as 0.5% by volume stimulated algal primary production and biomass in these
nitrogen-limited waters. Paerl (1985) and Paerl et al. (1990) did not estimate the proportion
of the total nitrogen inputs to these areas that entered as precipitation, but they do suggest
that algal blooms initiated by direct inputs of nitrogen from large rain storms could be
sustained by NO3" enriched runoff from nearby land masses. Terrestrial inputs of nitrogen
(from runoff) usually lag rainfall by 4-5 days in this region. These studies appear to be
unique in showing a direct link between nitrogen deposition and algal productivity, but do not
provide enough information to estimate the overall importance of deposition to the
maintenance of high algal biomass in these waters.
August 1991
10-188 DRAFT-DO NOT QUOTE OR CITE
-------
1 10.6.4.3 Evidence for Nitrogen Deposition Effects in Estuarine Systems—Case Studies
2 Complete nitrogen budgets, as well as information on nutrient limitation and seasonal
3 nutrient dynamics, have been compiled for two large estuaries, the Baltic Sea and the
4 Chesapeake Bay, and for the Mediterranean Sea. In the case of the Mediterranean, Loye-
5 Pilot et al. (1990) suggest that 50% of the nitrogen load originates as deposition falling
6 directly on the water surface. In the case of the Baltic and Chesapeake, deposition of
7 atmospheric nitrogen has been suggested as a major contributor to the eutrophication of the
8 estuaries (see below). Data for other coastal and estuarine systems are less complete, but
9 similarities between these two systems and other estuarine systems suggest that their results
10 may be more widely applicable. The discussion in this document is limited to these two
11 "case studies," with some speculation about how other estuaries may be related.
12 ...
13 The Baltic Sea
14 The Baltic Sea is perhaps the best-documented available case study of the effects of
15 nitrogen additions in causing estuarine eutrophication. Like many other coastal waters, the
16 Baltic Sea has experienced a rapidly increasing anthropogenic nutrient load; it has been
17 estimated that the supply of nitrogen has increased by a factor of 4, and phosphorus by a
18 factor of 8, since the beginning of the century (Larsson et al., 1985). The first observable
19 changes attributable to eutrophication of the Baltic were declines in the concentration of
20 dissolved oxygen in the 1960s (Rosenberg et al., 1990). Decreased dissolved oxygen
21 concentrations result when decomposition in deeper waters is enhanced by the increased
22 supply of sedimenting algal cells from the surface water layers to the sediments. In the case
23 of the Baltic, the spring algal blooms that now result from nutrient enrichment consist of
24 large, rapidly sedimenting algal cells, which supply large amounts of organic matter to the
25 sediments for decomposition (Enoksson et al., 1990). Since the 1960s, researchers in the
26 Baltic have documented increases in algal productivity, increased incidence of nuisance algal
27 blooms, and periodic failures and unpredictability in fish and Norway Lobster catches
28 (Fleischer and Stibe, 1989; Rosenberg etal., 1990).
29 It has now been shown by a number of methods that algal productivity in nearly all
30 areas of the Baltic Sea is limited by nitrogen. Nitrogen to phosphorus ratios range from
31 6:1 to 60:1 (Rosenberg et al., 1990), but the higher values are only found in the remote, and
August 1991
10-189 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
relatively unimpacted, area of the Bothnian Bay (between Sweden1 and Finland). Productivity
in the spring (the season of highest algal biomass) is fueled by nutrients supplied from deeper
waters during spring overturn (Graneli et al., 1990); deep waters are low in nitrogen and
high in phosphorus, resulting in N:P ratios near 5 (Rosenberg et al., 1990), suggesting
potential nitrogen limitation when deep waters are mixed with surface waters. Low nitrogen
to phosphorus ratios in deep water result from denitrification in the deep sediments (Shaffer
and Ronner, 1984). Primary productivity measurements in the Kattegat (the portion of the
Baltic between Denmark and Sweden) correlate closely with uptake of NO3", but not of PO43"
(Rydberg et al., 1990). Level n and IE nutrient enrichment experiments conducted in near-
shore areas of the Baltic, as well as in the Kattegat, indicate nitrogen limitation at most
seasons of the year (Graneli et al., 1990). Growth stimulation of algae has also been
produced by addition of rain water to experimental enclosures, in amounts as small as 10% of
the total volume (Graneli et al., 1990); rain water in the Baltic is enriched in nitrogen, but
phosphorus-poor. In portions of the Baltic where freshwater inputs keep the salinity low,
blooms of the nitrogen fixing blue green alga Aphanizomenon flos-aquae are common
(Graneli et al., 1990); blue green algal blooms are common features of nitrogen-limited
freshwater lakes (see Section 10.6.4.1), but are usually absent from marine waters.
Nitrogen budget estimates indicate that the Baltic Sea as a whole receives 7.3 X
1010 eq • yr'1 of nitrogen, of which 2.8 X 1010 eq • yr'1 (37%) comes directly from
atmospheric deposition (Rosenberg et al., 1990). Fleischer and Stibe (1989) report that the
nitrogen flux from agricultural watersheds feeding the Baltic have been decreasing since
ca. 1980, but that the nitrogen contribution from forested watersheds is increasing; they cite
both increases in nitrogen deposition, and the spread of modern forestry practices, as causes
for the increase. It should be noted, however, that the Baltic also experiences a substantial
phosphorus load from agricultural and urban lands, and that phosphorus inputs may help to
maintain nitrogen-limited conditions (Graneli et al., 1990). If the Baltic had received
consistent nitrogen additions (e.g., from the atmosphere or from agricultural runoff) in the
absence of phosphorus additions, it might well have evolved into a phosphorus-limited system
some time ago.
The physical structure of the Baltic Sea, with a shallow sill limiting exchange of water
with the North Sea (see the definition of a fjord, above) contributes to the eutrophication of
August 1991
10-190 DRAFT-DO NOT QUOTE OR CITE
-------
11 the basin, by trapping nutrients in the basin once they reach the1 deeper waters. Because the
2 larger algal cells that result from nutrient enrichment in the basin provide more nutrients to
3 the deep water through sedimentation, and because only shallow Waters have the ability to
4 exchange with the North Sea, it is estimated that less than 10% of nutrients added to the
5 Baltic are exported over the sill to the North Sea (Wulff et al., 1990). Throughout much of
6 the year (i.e., especially during the dry months) productivity in the Baltic is maintained by
7 nutrients recycled within the water column (Enbksson et al., 1990). The trapping of nutrients
8 within' the basin, and recycling of nutrients from deeper waters by circulation patterns,
9 '- suggest :that eutrophication of the Baltic is a self-accelerating process (Enoksson et al., 1990),
10 with a long time lag between reductions of inputs and improvements in water quality.
n • -.- -- . :. . - .•-•-...-:• ' -- .•'-,..•-•.•, '. -•
12 The Chesapeake Bay
13 The most complete attempts to estimate the relative importance of atmospheric
14 deposition to the overall nitrogen budget of an estuary or coastal ecosystem in the United
15 States were completed for the Chesapeake Bay by the mvironmental Defense Fund (EDF;
16 Fisher et al., 1988a), and by Versar, Inc. (Tyler,-1988) in 1988. Neither of these reports has
17 been published in a peer-reviewed arena, but the issue of atmospheric contributions to the
18 eutrophication of the Chesapeake has been widely discussed (and criticized) particularly after
19 the publication of the EDF report, and bears close examination for these reasons.
20 Both reports conclude that atmospheric deposition makes a substantial contribution
21 (25-40% of total inputs) to the nitrogen budget of the Chesapeake Bay. In both cases,
22 nitrogen budgets for the Bay were constructed via a number of steps, each of which involved
23 simplifying assumptions which bear further examination. Both reports calculate inputs from
24 atmospheric deposition to the Bay itself (Step #1), -atmospheric deposition to the watershed
25 (#2), fertilizer application in the watershed (#3), generation of animal wastes in the watershed
26 (#4), inputs from urban land use (#5), and point source inputs (#6). Once the total inputs to
27 the watershed and Bay were estimated, both reports calculated the proportion of the inputs
28 that were retained by the watershed (Step #7), and the proportion that were retained within
29 the rivers and tributaries feeding the Bay (#8).
30 The two reports had different goals, which make their results difficult to compare. The
31 EDF report (Fisher et al., 1988a) estimated the proportions of both NO3~ and NH4+
August 1991
10-191 DRAFT-DO NOT QUOTE OR CITE
-------
1 deposition to the total nitrogen budget of the Chesapeake (including all forms of nitrogen, and
2 both baseflow and storm flows). The Versar report (Tyler, 1988), on the other hand,
3 estimated only contributions of NO3", because NH4+ does not result from the burning of
4 fossil fuels, and excluded baseflow contributions. In addition, the Versar report used a range
5 of values both for the watershed contributions made by each nitrogen source (deposition,
6 fertilizers, etc.) and for the fraction of the inputs retained by the watershed (transfer
7 coefficients). This results in a wide range of budget values for each of the sources, and for
8 the relative importance of NO3", deposition to the budget, which complicates any comparison
9 of the results of the two studies. None-the-less, the two reports used similar methods in
10 developing their budgets, and a combined discussion of the uncertainties involved in each of
11 the steps listed above is warranted.
12 The results for the two budgets are presented in Table 10-28. Since the publication of
13 these budgets, additional information on such issues as dry deposition, and retention of
14 nitrogen by forested watersheds, has become available. This new information has been
15 compiled to produce a third "refined" budget, which is also presented in Table 10-28. The
16 assumptions that were used to construct the refined budget are outlined in each of the
17 discussions of individual budgeting steps below. .
18 The major uncertainty involved in calculating direct inputs to the Chesapeake from
19 atmospheric deposition (Step #1, above) is estimation of the contribution of dry deposition
20 (see also Section 10.2). Both reports use actual deposition monitoring data (i.e., from
21 NADP/NTN) to estimate the nitrogen load from wet deposition, and then assume that the rate
22 of dry deposition of nitrogen in, the watershed is equal to the rate of wet deposition. As
23 discussed earlier (see section on nitrogen inputs), the measurement of dry deposition is a
24 much vexed issue, and most researchers make educated guesses of rates of dry deposition by
25 assuming that they are some fraction of wet deposition rates. The assumption that dry
26 deposition is equal to wet deposition is probably reasonable for areas directly adjacent to
27 emissions sources (Summers et al., 1986),but the ratio of dry deposition to the sum of wet
28 and dry deposition may fall as low as 0.2 in locations remote from sources. For example,
29 Barrie and Sirois (1986) estimated that dry deposition contributed 21 to 30% of total NO3"
August 1991
10-192 DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-28. THREE NITROGEN BUDGETS FOR THE CHESAPEAKE BAY
Source of Nitrogen.
EDF Budget Versar Budget
(eq X 109 • yr-1) (eq X 109 • yf})
Refined Budget
(eq X 109 • yr-1)
Direct Deposition
N03- 0.6
NH4+ 0.3
N Load to Bay (from direct deposition)13 0.9
Forests ...
NO3' Deposition 6.4
NH4+Deposition 3.5 80%
Watershed Retention . 0.6 50%
In-stream Retention 1-0
Atmospheric NO3' Load to Bay (from forests)
N Load to Bay (from forests)13
Pasture Land
NO3'Deposition , .1.7
NH4+ Deposition 0.9
, Animal Wastes • 13.9
Watershed Retention 0.5
In-stream Retention 1.1
Atmospheric NO3" Load to Bay (from pastures)
N Load to Bay (from pastures)6
Cropland , ,
NO3" Deposition 1-8 }7Q%
NH4+Deposition .1.0..
Fertilizers 11-3
Watershed Retention 0.6
In-stream Retention . 4.2
Atmospheric NO3" Load to Bay (from cropland)
N Load to Bay (from cropland)13. • '
Residential/Urban
0.5
_a
0.5
6.0
_a
0.15
0.15
95%
50%
0.4
0.2
0.6
4.6
2.5
0.5
0.7
.1.2 , . 0.9
95%° -a 94-99% ' 0.5
50%c 8.4 50% 13.9
0.01-0.04 0.09
, 0.05-0.3 • 0.6
2.0 1.5
-*• 76-99% 0.8
5.9-19.3 50% 11.3
0.010.2 • 0.05
0.04r2.6 . 0.4
84.6%
35%
95 %d
35%
95%
35%
NO3' Deposition
NH4+ Deposition
Watershed Retention
In-stream Retention „ . •
Atmospheric NO3" Load to Bay (from urban areas)
N Load to Bay (from urban areas)13
Point Sources
NO3- LOAD TO BAY (FROM DEPOSITION)
TOTAL NITROGEN LOAD TO BAYb,
% of N from NO3" deposition
0.3
, 0.2
0.2
0.3
2.4
2.50
9.95
25%
0.5
35% -* .62-96%
0% 0.01-0.1 20%
, ; 0.01-0.1 ,
1.4-2,3
0.67-1.06
2.16-5.90
18-31%e
0.4
0.2
0.1
0.2
2.4
1.09
4.87
22.5%
50%
35%
,aThe Versar Budget (Tyler, 1988) does not calculate loads of NH4+. .
bFor the EDF Budget (Fisher et al., 1988a) and refined budget total nitrogen load to the Bay includes both NO3"
and NH4+. The Versar Budget (Tyler, 1988) includes only NO3'. .
•Watershed and In-stream retention values for pastureland in the EDF Budget apply only to animal wastes. For
atmospheric deposition, the cropland retention value (70%) was used.
d95% retention was used for animal wastes; 85% retention was used for deposition (see text).
The range of contributions of NO3" deposition to the total budget were calculated by comparing maximum to
maximum estimates, and minimum to minimum estimates. These combinations are more likely to occur during
extreme (e.g., very wet or very dry) years.
August 1991
10-193 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
deposition in eastern Canada. Baker (1991) concludes that dry deposition of NO3" is ca. 40%
of wet, while dry deposition of NH4+ is approximately 34% of wet (resulting in ratios of dry
to wet plus dry deposition of 0.29 and 0.25, respectively), for areas remote from emissions.
Li the most complete analysis of dry and wet deposition of NO3" to date, Sisterson et al. (in
press) report ratios of dry to wet plus dry deposition of 0.35 for two locations inside or near
the borders of the Chesapeake Bay watershed (State College, Pennsylvania, and West Point,
New York). Based on the results of these studies, it seems that the assumption made in the
two Chesapeake Bay nitrogen budgets (i.e., that dry deposition is equal to, wet) probably
overestimates the importance of dry deposition. The 0.35 ratio is used in constructing the
refined budget in Table 10-28.
The two reports (Fisher et al., 1988a; Tyler, 1988) also present different values for the
direct contribution of wet deposition to the Bay, because they use different methods to
estimate the spatial pattern of deposition in the Bay and its watershed. The EDF report uses
wet deposition values from the nearest NADP collector; the Versar report extrapolates
deposition values from isopleth maps of NO3" deposition. In addition, the Versar report
includes direct atmospheric inputs to,the tributaries of the Bay, as well as to the Bay itself
(Table 10-28). Aside from .problems with estimating dry deposition, it seems likely, that the
approach used in the Versar report for estimating deposition is more precise than that used in
the EDF report. The Versar values for wet deposition were therefore used in the,refined
budget, after adjusting them to reflect a 35% contribution from dry deposition. Ammonium
deposition was calculated for the refined budget by applying the ratio of NH4+ to NO3"
deposition reported in the EDF report to the estimated NO3" deposition values from the
Versar report (i.e., these values assume that the spatial,pattern in NH4+ deposition is the
same as the spatial pattern for NO3" deposition).
The uncertainties involved in estimating nitrogen deposition to the Chesapeake Bay
watershed (Step #2) are similar to those for estimating direct deposition. It seems likely that,
by assuming dry deposition is equal to wet, both reports overestimate the dryfall contribution
to deposition. Differences between the estimates of wet deposition presented in the two
reports result from the same methodological differences used in estimating direct inputs (i.e.,
use of the nearest NADP collector, vs. extrapolated values from isopleth maps), and slight ,
differences in the estimates of the coverage of each land use type. The Versar method
August 1991
10-194 DRAFT-DO NOT QUOTE OR CITE
-------
1 produces slightly lower estimates of atmospheric nitrogen inputs to the basin (Table 10-28)
2 and, as in the case with estimates of direct deposition to the Bay, the Versar method probably
3 produces better estimates of basin-wide deposition loads than the EDF approach. The refined
4 budget uses the Versar values for wet NO3" deposition (adjusted to reflect a 0.35 ratio for dry
5 deposition, as above), and estimates of NH4+ deposition based on the Versar spatial
6 deposition pattern, and the EDF estimate of NH4+deposition, as above.
7 The EDF report (Fisher et'al., 1988a) uses county agricultural reports and U.S. Census
8 Bureau data to calculate the application rates of fertilizers to the counties (and portions
9 thereof) in the Chesapeake Bay watershed (Step #3, above): The Versar report (Tyler, 1988)
10 calculates the total fertilizer load (from NO3") to the watershed by applying a correction
11 factor to the level of fertilizer application recommended by the U.S. Department of
12 Agriculture; the correction factor was based on local officials' best guesses of actual fertilizer
13 application rates (e.g., 30 to 60% of the recommended rates). Because it deals only with
14 NO3" loading, the Versar approach also necessitates making an assumption about the
15 proportion of nitrogen fertilizers that are applied as NO3", as opposed to NH4+ or urea; the
16 report assumes that 60% of the nitrogen added is in the form of NO3", but presents no data to
17 ' support this assumption. Because it is more direct in nature, the EDF approach to estimating
18 fertilizer inputs seems to be more defensible than the Versar approach, and the EDF estimate
19 is therefore used in the refined budget. The EDF estimate of 11.3 X 109 eq • yr'1 is near
20 the bottom range of fertilizer loads estimated by the Versar report (Table 10-28).
21 The EDF (Fisher et al., 1988a) and Versar (Tyler, 1988) reports use the same estimate
22 (from the EDF report) for the contribution by animal wastes (Step #4, above) to the nitrogen
23 budget. The EDF report used county agricultural statistics to calculate the total number of
24 farm animals of different types in the Chesapeake Bay watershed. These population numbers
25 were then multiplied by published estimates of the amount of nitrogenous wastes excreted by
26 each type of animal annually, to produce an estimate' of 13.9 x 109 eq • yr"1. As in the
27 estimates of fertilizer NO3" inputs, the Versar report assumed that 60% of animal nitrogenous
28 wastes were in the form of NO3"; this estimate seem especially difficult to justify when it is
29 used both for animal wastes and for fertilizers, as there is no reason to expect both nitrogen
30 sources to have the same composition. The EDF estimate of 13.9 X 109 eq • yr"1 is used for
31 the refined budget.
August 1991
10-195 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
In both reports, atmospheric deposition is considered to be the only source of nitrogen
to urban areas (Step #5, above). As pointed out in the Versar report (Tyler, 1988), this is
likely to be an underestimate because it ignores fertilizer applications to lawns and gardens.
Because fertilizers applications are seasonal, and the area of urban land in the basin is small
(about 3% of the total), this underestimate is considered unimportant. As mentioned earlier,
the EDF (Fisher et al., 1988a) and Versar reports use slightly different methods to calculate
wet deposition. The primary difference between the two estimates of nitrogen loading to
urban areas (Table 10-28), however, is in their estimate of the proportion of the basin in
residential and urban land use (5 x 105 ha in the EDF report vs. 8 x 105 ha in the Versar
report). In neither case does the nitrogen contribution from urban lands (<2% of the total
loading to the watershed) play a significant role in the budgets. The Versar estimate of
deposition to urban areas is used in the refined budget, with the same adjustments applied as
for the deposition to the watershed/and directly to the Bay (above).
Both reports used the same EPA estimates of point source inputs to the Chesapeake Bay
watershed (Step #6, above); the lower value presented in the Versar report (Tyler, 1988) is
the estimated proportion of point source inputs that are in the form of NO3", again assuming
that NO3" is 60% of total inorganic nitrogen. The upper limit to the range of point source
inputs presented by the Versar report is a more recent (1988) estimate from the Chesapeake
Bay Program. There seems to be little reason not to use the original EDF value (Fisher
et al., 1988a) of 2.35 X 109 eq • yr'1 (Table 10-28), and this value is used in the refined
budget.
Perhaps the greatest source of uncertainty in both nitrogen budgets is created when the
proportions of nitrogen inputs that are retained within the watershed are estimated (Step #7,
above). Both reports use a variety of methods to calculate separate transfer coefficients for
each land use type, and in some cases, for different sources of nitrogen within a single land
use type. In particular, the Versar report (Tyler, 1988) compares calculated loads (as
described in the preceding paragraphs) to calculated runoff from each land use type (from
Smullen et al., 1982) and estimates a range of transfer coefficients from these calculated
values. Because the error inherent in the calculated values is amplified when they are
compared, this method seems especially problematic. Often the calculated transfer
coefficients differ greatly from coefficients measured for single basins within the Chesapeake
August 1991
10-196 DRAFT-DO NOT QUOTE OR CITE
-------
1 Bay watershed. The transfer coefficients for each land use type are discussed in detail below.
2 It should be emphasized that all of the nitrogen budgets discussed below deal only with
3 inorganic forms of nitrogen (i.e., NO3" and NH4+). Outputs of organic nitrogen from
4 watershed can be substantial (e.g., Correll and Ford, 1982), and organic forms can result
5 from atmospheric deposition sources when watershed processes route nitrogen through the
6 biotic portion of the ecosystem. Given this possible source of error, the nitrogen retention
7 values presented below should probably be considered maximum estimates.
8 Estimating watershed retention of nitrogen in forested watersheds is difficult, primarily
9 because so few data are available, and the applicability of single watershed values to wide
10 areas of the Chesapeake Bay watershed is untested. The Versar (Tyler, 1988) report
11 compares calculated deposition loads (Table 10-28) to estimates of runoff from forests (from
12 Smullen et al., 1982) to yield a transfer coefficient of 4.8%. As discussed above, this
13 estimate must be considered very uncertain, because of the combined errors introduced by
14 comparing two calculated values. The EDF report (Fisher et al., 1988a) found literature
15 values that ranged from 50% (in the Mid-Appalachians) to 97% (in the Coastal Plain), and
16 used 80% as a "reasonable mid-range estimate." Given the range of possible retention
17 values, it seems unlikely that any single number would be a reasonable estimate for the entire
18 Chesapeake Bay watershed. Some additional nitrogen retention values are given in
19 Table 10-29, based on published nitrogen budgets for watersheds in or near the Chesapeake
20 Bay basin. These are arranged according to physiographic regions, in order to illustrate the
21 spatial variability in watershed nitrogen retention. Of the values in Table 10-29, only those
22 of Kaufmann et al. (1991) are applicable to broad spatial areas, because they are based on a
23 probability sampling of streams in each region. These values assume that NO3"
24 concentrations at spring baseflow are representative of annual mean concentrations (Kaufmann
25 et al., 1988; Messer et al., 1988). If the retention coefficients for each physiographic region
26 are weighted by the proportion of the Chesapeake Bay watershed in each physiographic
27 region (from Smullen et al., 1982), an area-weighted retention coefficient of 84.6% results;
28 this figure was used for the refined budget (Table 10-28). The 84.6% figure agrees
29 remarkably well with the data presented in Figure 10-16b (Driscoll et al., 1989a), which
30 suggest an interpolated coefficient of 84.7% at the levels of deposition calculated for the
August 1991
10-197
DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-29. WATERSHED RETENTION OF NITROGEN IN WATERSHEDS
IN OR NEAR THE CHESAPEAKE BAY BASIN, FROM PUBLISHED REPORTS.
ALL NITROGEN LOADS HAVE BEEN RE-ESTIMATED BASED ON MEASURED
WET DEPOSITION, AND A 35% CONTRIBUTION TO TOTAL DEPOSITION
FROM DRY DEPOSITION
Physiographic Region
Poconos/Catskills"
Biscuit Brook, NY
Northern Appalachiansb
Southwestern Pennsylvania
Southwestern Pennsylvania1"
Fernow, WV
Eastern Tennessee
Valley and Ridgeb
Catoctin Mountains, MD
Shenandoah National Park, VA
Mid-Atlantic Coastal Plainb
Chesapeake Bay, MD
Piedmont*
Northern Georgia
Southern Blue Ridgeb
eq X
N Load'
-
878
-
1,192
-
1,506
707
-
593
557
-
1,000
-
486
'
106 • yr'1
NO3" Export
-
214
-
264
-
607
36
-
250
3
-
10
. -
11
-
% Retention
88.3%
75.7%
72.7%
78.0%
94.5%
59.5%
94.6%
78.5%
57.5%
99.5%
90.9%.
99.0%
90.2%
97.7%
88.3%
Source
Kaufmann et al. (1991)
Stoddard and Murdoch (1991)
Kaufmann et al. (1991)
Barker and Witt (1990)
Sharpe et al. (1984)
Helvey and Kunkle (1986)
Kelly (1988)
Kaufmann et al. (1991)
Katz et al. (1985)
Shaffer and Galloway (1982)
Kaufmann et al. (1991)
Weller et al. (1986)
Kaufmann et al. (1991)
Buell and Peters (1988)
Kaufmann et al. (1991)
•Nitrogen loads are calculated from published wet deposition estimates, extrapolated to total deposition
according to a 0.35 dry:wet plus dry ratio (see text)
^Retention estimates are calculated by comparing mean concentrations of precipitation to mean concentrations i
stream water. Estimates from Kaufmann et al. (1991) are from the National Stream Survey (Kaufmann
et al., 1988) and are for the population of streams within each physiographic province.
in
1
2
3
4
5
6
7
8
9
10
Chesapeake Bay watershed (636 eq • ha"1 total deposition, or 413 eq • ha"1 of wet
deposition).
Nitrogen retention by pasturelands is generally thought to be very high. Both the EDF
(Fisher et al., 1988a) and Versar (Tyler, 1988) reports estimate retention coefficients in the
94-99% range. As with forest nitrogen retention, the EDF estimate is based on published
values from watershed studies, while the Versar estimate is based on comparisons of
calculated loads and calculated runoff. The EDF estimate (95%) is based primarily on a
study by Kuenzler and Craig (1986; as reported in Fisher et al., 1988a) on pastureland in the
Chowan River watershed. Similar results (94.4% retention) have been reported for
unfertilized pasture lands in Ohio by Owens et al. (1989), where NO3" losses were lower
August 1991
10-198 DRAFT-DO NOT QUOTE OR CITE
-------
1 , from pastureland than from nearby undisturbed forests (86% retention). Nitrogen retention
2 coefficients reported here were recalculated to include dry deposition (at 35% of total
3 deposition), as was the case for forest nitrogen budgets reported above. The EDF report
4 applies the 95% retention rate only to animal wastes, and uses a 70% retention coefficient for
5 atmospheric deposition. Because they are primarily in the form of particulate organic matter,
6 it seems reasonable to assume that animal wastes will be more strongly retained than
7 deposition. The refined budget therefore applies the 95 % retention figure for animal wastes,
8 and an 85% retention coefficient (as for forests, above) for nitrogen from deposition
9 (Table 10-28).
10 The ability of croplands to retain nitrogen is generally high, because most of the
11 nitrogen applied to crops as fertilizer is removed as biomass during harvest (Lowrance et al.,
12 1985; Groffman et al., 1986). Both the EDF (Fisher et al., 1988a) and Versar (Tyler, 1988)
13 budgets compare estimates of fertilizer and deposition loads to estimates of runoff from
14 croplands to calculate nitrogen transfer coefficients. Use of loads estimates from a number of
15 sources creates a range of retention coefficients from 70% (Fisher et al., 1988a) to 99%
16 (Tyler, 1988). Published values from studies of cropland watersheds are all toward the
17 higher end of this range. Peterjohn and Correll (1984) measured a retention coefficient of
18 '93.2% for a fertilized corn field in Maryland. Groffman et al. (1986) report 100% retention
19 of fertilizer nitrogen in a sorghum field in the Georgia piedmont; lower retention coefficients
20 (76.1 %) were measured during the winter, but the planting of crimson clover (a nitrogen-
21 fixing legume) as a winter cover crop complicates the interpretation of these figures.
22 Lowrance et al. (1985) report nitrogen budgets for four cropland watersheds with a variety of
23 crops in the Georgia Coastal Plain, with retention coefficients ranging from 97.8 to 100%.
24 Nitrogen retention coefficients reported here were recalculated to include dry deposition (at
25 35% of total deposition), as was the case for forest and pastureland nitrogen budgets reported
26 above. A retention coefficient of 95 %, as used for the refined budget (Table 10-28) is near
27 the middle of the range of published values. Fertilizer inputs are generally in the same
28 inorganic forms as atmospheric deposition, and there seems no reason to apply different
29 retention values to fertilizer and deposition sources of nitrogen.
30 Published reports of nitrogen retention in urban lands are apparently unavailable. The
31 EDF report (Fisher et al., 1988a) simply chose a retention coefficient midway between their
August 1991
10-199
DRAFT-DO NOT QUOTE OR CITE
-------
1 cropland value (70%) and complete runoff from impervious surfaces (100%). The Versar
2 report (Tyler, 1988) calculates transfer coefficients from estimated loads (from deposition)
3 and estimated runoff, and gives a range of 62% to 96% (Table 10-28). There is little
4 justification for choosing any particular value. The 50% value used for the refined budget
5 (Table 10-28) is chosen only to provide a "ball-park" value; slightly higher or lower values,
6 when applied to the relatively small atmospheric loads falling on urban areas, will not
7 substantially change the conclusions presented here.
8 The final assumption that affects the nitrogen budgets concerns the proportion of
9 watershed runoff that is lost during transport through rivers to the Bay (step #8, above).
10 Denitrification in slow moving lotic waters can significantly reduce the load of nitrogen
11 delivered to estuarine waters (see Section 10.6.2.4). In the absence of any measured loss
12 rates, both the EDF (Fisher et al., 1988a) and Versar (Tyler, 1988) reports adopt the 50%
13 loss value suggested by the Chesapeake Bay Program (Smullen et al., 1982). More recently,
14 denitrification values have been published for two rivers, the Potomac, which supplies water
15 directly to the Chesapeake Bay, and the Delaware, which is adjacent to the Chesapeake Bay
16 watershed (summarized in Seitzinger, 1988a). Seitzinger (1987) estimated that 35% of the
17 dissolved inorganic nitrogen (NO3~ + NH4+) load to the Potomac River was lost through
18 denitrification. Seitzinger (1988b) measured denitrification rates at six locations in the tidal
19 portion of the Delaware River, and estimated that 20% of the dissolved inorganic nitrogen
20 load was lost through denitrification. Both of these studies were conducted in the relatively
21 flat, slow-moving and tidal portions of rivers, where denitrification rates are likely to be
22 maximal, due to the existence of anoxic sediments. Data from smaller streams suggest that
23 lower rates of nitrogen retention (10-15%) are more likely to occur in headwater streams
24 (Triska et al., 1990; Duff and Triska, 1990). In light of these lower measured rates of
25 nitrogen loss, the 50% figure used in the EDF and Versar budgets seems insupportable for
26 riverine losses; loss rates as high as 50% have been measured only in estuarine waters (e.g.,
27 Narragansett Bay, Seitzinger et al., 1984; Baltic Sea, Larsson et al., 1985). The refined
28 budget uses a figure of 35%, reflecting the only known value for a river feeding the
29 Chesapeake itself (Seitzinger, 1987), and may still overestimate in-stream retention in small
30 streams.
August 1991
10-200
DRAFT-DO NOT QUOTE OR CITE
-------
1 When the three budgets are compared, they suggest a wide range in estimated
2 contributions from individual sources of nitrogen (e.g., estimates of cropland inputs vary
3 from 0.04 X 109 eq • yr"1 for the "best case" Versar budget, to 4.2 X 109 eq • yr"1 for the
4 EDF budget), but a surprisingly consistent percentage contribution from atmospheric NO3"
5 deposition (18 - 31%) to the total budget (Table 10-28). All three budgets suggest that a
6 large amount of nitrogen enters the Bay from deposition; the 1.1 x 109 eq • yr"1 estimate
7 from the refined budget corresponds to a nitrogen load of 19 tons per day entering the
8 Chesapeake Bay from deposition directly to the Bay and the watershed. The caveat presented
•9 earlier concerning organic forms of nitrogen should probably be repeated here; the estimates
10 of atmospheric NO3" contributions to the Bay ignore all but the inorganic nitrogen fractions.
11 Organic nitrogen can be a substantial contributor to the nitrogen in runoff, and could
' 12 potentially have a large atmospheric deposition component. Many of-the estimates that went
13 into these budgets are relatively certain. For example, we have good data on wet deposition,
14 and can extrapolate to total deposition with reasonable certainty given recent estimates of dry
15 deposition within the watershed (e.g., Sisterson et al., in press). The biggest uncertainty in
16 estimating atmospheric NO3" loading to the Bay comes results from the figure for retention of
17 nitrogen by forested watersheds. This influence results from the fact that most of the
18 watershed (ca. 80%) is forested; small changes in the retention coefficients can have a large
19 effect on the estimated load to the Bay from these watersheds. The retention coefficient
20 calculated for the refined budget (84.6%) is our current best estimate, based on regional
21 estimates of retention within each of the physiographic regions in the Chesapeake Bay basin,
22 however it still contains considerable uncertainty. The retention coefficients listed in
23 Table 10-29 suggest that retention can vary from less than 60% to more than 99% in
24 individual watersheds. Many more values from individual watersheds are needed before we
25 can be certain how representative the values for each physiographic region are.
26 Taken as a whole, the budgets suggest that deposition is approximately equal in •
27 importance to point source supplies of nitrogen, and is possibly more important than
28 agricultural sources of nitrogen (Table 10-28). The fact that three different approaches (i.e.,
29 the three budgets in Table 10-28) yield similar results lends weight to the suggestion that
30 atmospheric nitrogen contributes substantially to the eutrophication of the Chesapeake Bay.
31 These results are surprising, given the emphasis usually placed on reducing point source
August 1991
10-201
DRAFT-DO NOT QUOTE OR CITE
-------
1 inputs to the Bay in order to improve water quality (e.g., Chinchilli, 1989; Caton, 1989).
2 Based on the results of nutrient limitation work discussed earlier, it seems clear that the
3 control of nitrogen inputs is important to the control of eutrophication in the Chesapeake Bay.
4 The results of the budget exercises discussed here suggest that any program for nitrogen
5 control should include the control of nitrogen deposition, as well as point and non-point
6 sources.
7 Finally, atmospheric NO3" inputs to the Chesapeake should be put into the context of
8 seasonal nitrogen limitation of algal productivity in the Bay. As was discussed earlier, the
9 Bay may undergo seasonal shifts hi nutrient limitation, from phosphorus limitation in late
10 winter and early spring, to nitrogen limitation during summer and fall (e.g., D'Elia et al.,
11 1982; D'Elia et al., 1986). If atmospheric NO3" is to have a significant effect on algal
12 biomass, it would need to be present during the late summer, low flow, high biomass period.
13 However, much of the NO3" load occurs during the spring, when river flows and NO3"
14 leakage from watersheds are high (e.g., Lowrance and Leonard, 1988). In the case of the
15 Baltic Sea, discussed earlier, nutrients were largely trapped within the estuary by
16 sedimentation processes, and minimal water exchange with the North Sea. Does the
17 Chesapeake Bay act in a similar manner to trap nutrients, providing a mechanism for
18 springtime loads of NO3" to influence summertime productivity? Unfortunately, few
19 measurements of the nutrient retention capacities of the Chesapeake are available, but some
20 estimates have been made. Smullen et al. (1982) estimated, based on some measurements of
21 current and nutrient concentrations at the mouth of the Bay, and a simple box model, that
22 virtually all of the nitrogen entering the Bay was retained. Nixon (1987) and Nixon et al.
23 (1986) question this conclusion, and point out the such high nutrient retention rates should
24 result in very high nutrient concentrations in the sediments, which have not been found.
25 Based on estimates of sediment nutrient concentrations, Nixon et al. (1986) calculate that only
26 ca. 5% of nitrogen entering the Bay is retained. The argument of Nixon et al. (1986),
27 however, seems to ignore the potential effect of denitrification in maintaining low sediment
28 nitrogen concentrations, despite high rates of retention by the Bay. Fisher et al. (1988b) use
29 longitudinal profiles of nutrient concentrations through the Bay to estimate that 33 to 71% of
30 nitrogen entering the bay is retained. These lower estimates of nitrogen retention suggest that
31 nitrogen entering the Bay during spring runoff does have the potential to affect productivity
August 1991
10-202 DRAFT-DO NOT QUOTE OR CITE
-------
1 in the Bay during the critical summer months. They also suggest, however, that the
2 Chesapeake could return to background nitrogen concentrations within several flushing times
3 of the Bay, or within several years (Fisher et al., 1988b), if nutrient control strategies were
4 put in place.
5 It is impossible to determine at this point whether the Chesapeake Bay example is an
6 unusual one in terms of the relative importance of atmospheric nitrogen inputs. Jaworski
7 (1981) gives crude nitrogen budgets for four estuaries and embayments in the United States;
8 his results suggest that the Chesapeake Bay receives an unusually large proportion of nitrogen
9 (68%) from land runoff (which includes agricultural and deposition sources). Jaworski's
10 (1981) budgets indicate that wastewater discharges are more important in the Hudson River
11 (New York) and San Joaquin River (California) estuaries (63% and 47% of inputs,
12 respectively, but these estimates do not include deposition), while the Potomac River estuary
13 has equal inputs from wastewater and land runoff. Of Jaworski's four systems, the
14 Chesapeake Bay is the least influenced by point source pollution, but it also receives larger
15 inputs from point sources than many estuaries in the United States (e.g., the Apalachicola
16 Bay, Nixon and Pilson, 1983). If one views all estuarine and coastal waters as lying along a
17 gradient from high- to low-influence by point source pollution, then the relative importance
18 of deposition to the nitrogen budget will change as one moves along the gradient. The
19 general applicability of the nitrogen budget results from the Chesapeake Bay will depend on
20 where the Bay falls along this gradient.
21
22 10.6.5 Direct Toxicity Due to Nitrogen Deposition
23 In addition to the effects of acidification and eutrophication, nitrogen deposition could
24 potentially contribute to directly toxic effects in surface waters. Toxic effects on freshwater
25 biota result from un-ionized ammonia (NH3) that occurs in equilibrium with ionized
26 ammonium (NH4+) and OH". High NH3 concentrations are associated with lesions in gill
27 tissue, reduced growth rates of trout fry, reduced fecundity (number of eggs), increased egg
28 mortality, and increased susceptibility of fish to other diseases as well as a variety of
29 pathological effects in invertebrates and aquatic plants (reviewed in U.S. Environmental
30 Protection Agency, 1985). Most analytical methods for ammonium actually measure the sum
31 of ammonia (NH3) and ammonium (NH4+), which is commonly referred to as NH4+; for
August 1991
10-203
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
S
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
clarity, the sum of ammonium and ammonia will be referred to here as total ammonia
CT-NH3). No single toxic concentration for T-NH3 can be established, because the relative
contribution of NH3 to T-NH3, and the toxicity of NH3, vary with the pH and temperature
(Emerson et aL, 1975), and the ionic strength (Messer et al., 1984) of the water. The
proportion of NH3 increases at higher temperatures and increasing pH. Because of the
variability in NH3 toxicity, hew criteria have recently been developed which calculate the ..
toxicity as a function of pH, temperature and ionic strength (U.S. Environmental Protection
Agency, 1985). The new regulations require the calculation of a "final chronic value" (FCV)
and "final acute value" (FAV); 4-day average concentrations of NH3 cannot exceed the FCV
more often on average than once every three years, nor can 1-h average concentrations
exceed one-half of the FAV more often on average than once every three years.
Critical concentrations of NH3 that cause the various effects are wide ranging and are
related to site specific temperature and pH values. For example, 48-h LC50 values (the
concentration at which 50% of the test organisms die within 48 h) for Daphnia magna, a
common invertebrate found in lake zooplankton, range from 38 to 350 jumol • L"1 T-NH3
over a temperature range from 19.6 °C to 25 °C and pH range of 7.4 to 8.6 (U.S.
Environmental Protection Agency, 1985). However, results of toxicity tests on stream insects
showed that 96-h LC50 values ranged from 128 to 421 /zmol • L"1 T-NH3 at relatively
constant chemical conditions. The 96-h LC50 values for rainbow trout range from 11.4 to
78.5 #mol • L"1 T-NH3. Fingerlings tend to be less sensitive than older life stages and lower
oxygen concentrations increased sensitivity to NH3. Variation in temperature, pH,
acclimation time, and CO2 concentrations also appeared to explain some variation in
responses. Effler et al. (1990) calculate FCV and FAV values for Onondaga Lake, an urban
lake in Syracuse, New York, that is heavily polluted with municipal sewage. For both
salmonid and non-salmonid fishes, the FCV values varied (with time of year) from 1.4 to
2.9 /Jmol • L"1. One-half FAV values for non-salmonids varied from 3.6 to 28.6 /Jmol • L"1
(acute toxicity information for salmonids is not given). At typical pH (pH = 8) and
temperature (temperature = 20 °C) values for Onondaga Lake, the minimum FCV value of
1.4 jamol • L"1 corresponds to a T-NH3 concentration of 36 /zmol • L"1; this concentration is
always exceeded in the lake (Effler et al., 1990).
August 1991
10-204
DRAFT-DO NOT QUOTE OR CITE
-------
1 Onondaga Lake is unusual in being very productive, and so tends to be warmer and
2 have a higher pH than many lakes. At lower pH values (pH = 7) and lower temperatures
3 (15 °C) the percentage of T-NH3 that is free NH3 drops dramatically (Emerson et al., 1975),
4 so that the FCV values reported for Onondaga Lake would not be exceeded until a T-NH3
5 concentration of 785 #mol • L"1 was reached. Currently no areas of North America are
6 ~ known to experience rates of NH4+ deposition that are sufficient to produce such high
7 concentrations in surface waters. Given current maximal concentrations of NH4+ in
8 deposition (40 /zmol L"1, Stensland et al., 1986) and reasonable maximum rates of dry
9 deposition and evapotranspiration (dry equal to 100% of wet deposition, and
10 evapotranspiration equal to 50% of deposition), maximum NH4+ concentrations in surface
11 waters will be less than 160 /rniol • L"1. If all nitrogen in deposition (NO3~ + NH4+) were
12 ammonified, maximum potential NH4+ concentrations attributable to deposition would be
13 approximately 280 #mol • L"1, and unlikely to be toxic except in unusual circumstances.
14 Since NH4+ is rapidly oxidized to NO3" in watershed soils, and under well oxygenated
15 conditions in lakes and streams, the likelihood of reaching toxic concentrations are extremely
16 limited. Toxic levels would be more likely in systems that have oxygen deficits, high organic
17 matter loading (which would increase oxygen demand and contribute ammonium through
18 mineralization processes), and direct inputs of ammonia (i.e., near feed lot operations). In
19 such cases it would probably be more effective to remove the local causes of oxygen
20 depletion, and organic matter loading, than to reduce atmospheric inputs of nitrogen. It
21 appears from the information above that the potential for directly toxic effects attributable to
22 nitrogen deposition in the United States is very limited.
23 .
24
25 10.7 DISCUSSION AND SUMMARY
26 10.7.1 Introduction
27 Since the mid-1980s the view has emerged that the atmospheric deposition of inorganic
28 nitrogen has impacted aquatic and terrestrial ecosystems. In many areas of the United States
29 it is known that the atmospheric input of nitrogen compounds has been significant (U.S.
30 Environmental Protection Agency, 1982, Section 10.7.2), however, the impacts have
31 generally been unknown or considered benign. Although, the evidence linking nitrogen
August 1991
10-205
DRAFT-DO NOT QUOTE OR CITE
-------
1 deposition with ecological impacts has been tenuous, there has been a growing concern
2 (Skeffington and Wilson, 1988). This concern has been magnified because (1) the
3 atmospheric concentrations of nitrogen compounds have increased in North America and most
4 European countries and (2) ecosystems formerly limited by nitrogen have become nitrogen
5 saturated via atmospheric deposition. These concerns have led to the efforts to develop
6 "critical loads" of nitrogen for various ecosystems. A "critical load" is defined as,
7 "a quantitative estimate of an exposure to one or more pollutants below which significant
8 harmful effects on specified sensitive elements of the environment do not occur according to
9 present knowledge" (Nilsson and Grennfelt, 1988).
10 Human welfare is dependent on ecological systems and processes. Natural ecosystems
11 are traditionally spoken of in terms of their structure and functions. Ecosystem structure
12 includes the species (richness and abundance), their mass and arrangement in an ecosystem.
13 This is an ecosystem's standing stock—nature's free "goods" (Westman, 1977). Society reaps
14 two kinds of benefits from the structural aspects of an ecosystem: (1) products with market
15 value such as fish, minerals, forest and pharmaceutical products, and genetic resources of
16 valuable species (e.g., plants for crops, timber, and animals for domestication) and (2) the
17 use and appreciation of ecosystems for recreation, aesthetic, enjoyment and study (Westman,
18 1977). «
19 Ecosystems have both structure and function. The most visible levels of organization
20 are: (1) the individual and its environment; (2) the population and its environment; and
21 (3) the biological communty and its environment, the ecosystem (Billings, 1978).
22 Ecosystems function as energy and nutrient transfer systems. Vegetation through the process
23 of photosynthesis accumulates, uses, and stores carbon compounds (energy), to maintain their
24 physiological processes amd to build plant structure. Carbohydrates and other compounds
25 accumluated and stored by plants are the basic source of food (energy and nutrients) for the
26 majority of animals and microorganisms. Energy moves unidirectionally and ultimately
27 dissipates into the environment. Nutrients are .recycled into the system. Because the various
28 ecosystem components are chemically interrelated, stresses placed on individual components,
29 such as those caused by nitrogen deposition and loading, can produce perturbations that are
30 not readily reversed and will significantly alter the ecosystem.
31
August 1991
10-206 DRAFT-DO NOT QUOTE OR CITE
-------
1 10.7.2 The Nitrogen Cycle
2 Nitrogen, one of the main constituents of the protein molecules essential to all life, is
3 , recycled within ecosystems. Most organisms cannot use the molecular nitrogen found in the
4 earth's atmosphere. It must transformed by terrestrial and aquatic microorganisms into a
5 form other organisms can use. The. transformations of nitrogen as it moves through the
6 ecosystem is referred to as the ."nitrogen cycle." Mature natural-ecosystems are essentially
7 self-sufficient and independent of external additions. Modern technology by either adding
8 nitrogen or removing nitrogen from ecosystems may be upsetting the relationships that exist
9 among the various components and thus changing its structure and functioning.
10
11 10.7.3 Nitrogen Deposition
12 The removal (dry deposition) of reactive nitrogen gases from the atmosphere occurs
13 along several pathways leading, to foliage, bark, or soil with pathways to foliage being
14 predominant during the growing season. The prevalence of any particular type of deposition
15 is a function of (1) the physicochemical properties of nitrogen compounds, (2) their ambient
16 concentration, and (3) the presence of suitable receptor sites in the landscape (e.g., leaves
17 with open stomata). Average canopy-level measurements (Table 10-30) exhibit the following
18 pattern or tendency towards dry deposition HNO3 > NH3 = NO2 > NO. Although the
19 leaf-level data for crops is incomplete (NO and HNO3 data are not available), the leaf
20 conductance (Kx) data for trees shows a similar pattern. These patterns are consistent with
21 observations of Bennett and Hill (1973), and can be partially explained by gas solubility
22 characteristics (Taylor et al., 1988). Particle deposition data averaged across species and
23 experimental techniques shows approximately three times greater nitrate aerosol deposition
24 (7.8 mm s'1) than for ammonium (2 mm s'1). However, the high average depositon velocity
25 (Vd) for NO3' is probably excessively high due to the unavoidable inclusion of nitrate from
26 HNO3 in measurements of nitrate deposition.
27 With the possible exception, of nitric acid vapor, deposition characteristics of reactive
28 nitrogen compounds are highly variable and dramatically influenced by environmental
29 conditions that effect .stomatal conductance. The tight relationship between stomatal
30 conductance and the deposition of NO and NO2 implies that gaseous deposition of reactive
31 nitrogen oxides is greatly reduced in the dark when stomata close (Hanson et al., 1989; Saxe,
August 1991
10-207 DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-30. MEAN DEPOSITION CHARACTERISTICS OF REACTIVE
NITROGEN GASES AT THE LEAF OR CANOPY SCALE OF
RESOLUTION FOR CROP OR TREE SPECIES
Compound
Summary for Crop Species
NO
N02
HN03
NH3
Summary for Tree Species
NO
NQ2
HNO3
NH3
Leaf Level Measures
Kj (mm s'1)3
• NDb ,
1.2
ND
.4.5
<0.3 '
1.1
2.1- •
1.8
Canopy Level Measures
Vd (mm s"1)a
1.3 ;
7.7
19.8
6.6
ND
24
41
22
'Means are the average for all species studied. However, measurements on dormant plant materials, foliage
with low stomatal conductance, and data recorded in the dark were excluded. The values listed as K, and Vd
for particles represent the leaf wash and throughfall measurement techniques, respectively.
bND; No data for crop plants were available.
1
2
3
4
5
6
7
8
9
10
1986; Hutchinson et al., 1972). Deposition of gaseous N forms is usually proportional to
ambient concentrations, but "compensation concentrations" at which no uptake occurs (i.e.,
<0.003-0.005 ppmv) have been reported for NO2 and NH3. Data for NO, NO2, and HNO3
(Grennfelt et al., 1983; Johansson, 1987; Marshall and Cadle 1989; Skarby et al., 1981),
from the vegetation dormant period, show a reduced potential for deposition. Conversely,
participate nitrate and ammonium deposition do not appear to be affected by the season of the
year (Gravenhorst et al., 1983; Lovett and Lindberg 1984).
The preceding information on gases and particles indicates that methods for measuring
gas or particulate deposition may produce dramatically different, results. Leaf-level measures
of deposition (Kj) for NO, NO2, and HNO3 were 4 to 10 times lower than estimates obtained
August 1991
10-208 DRAFT-DO NOT QUOTE OR CITE
-------
1 using micrometeorological canopy-level measurements (Vd). This discrepancy can largely be
2 explained once canopy area instead of ground area is factored into the canopy based
3 measurements. ..•'.••
4 The canopy-level deposition velocity measurement (Vd) has been criticized because it
5 attempts to pool environmental, physiological, and morphological characteristics into a single
6 descriptive measurement (i.e., it attempts to dp too much; Taylor et al., 1988). The result of
7 this over simplification is that Vd for even a single trace gas varies substantially in space and
8 time. However, average KL and Vd values for NH3 on crop species were comparable,
9 perhaps because crop canopies are more uniform and closer to the ground. Particle
10 deposition is governed by a different set of principles (see section 10.2.3) and the same
11 relationships between leaf and canopy level measurements may not be applicable.
12 Daytime rates of nitrogen oxide or ammonia deposition can also be approximated from
13 ambient concentrations of the gases (U.S. Environmental Protection Agency, 1982; Hicks
14 et al., 1985), and deposition constants such as those presented in Table 10-30. Hanson et al.
15 (1989) used such information with conservative estimates of concentration to approximate
16 total N deposition from nitrogen dioxide to various forest stands. They predicted NO2-N
17 inputs between 0.04 and 1.9 kg N ha"1 yr"1 for natural forests and inputs up to 12 kg N ha"1
18 yr"1 for forests in urban environments. For a forested watershed, Grennfelt and Hultberg
19 (1986) calculated the annual deposition of NO2 plus HNO3 to be in the range from 3.6 to
20 5.1 kg N ha"1 yr"1. Hill (1971) estimated the removal of NO2 from the atmosphere in
21 Southern California to be approximately 109 kg N ha"1 yr"1.
22 Preliminary particle deposition measurements and calculated dry deposition estimates of
23 reactive nitrogen gases, indicate significant nitrogen inputs to terrestrial systems. Barrie and
24 Sirois (1986) estimated that dry deposition contributed 21 to 30% of total NO3" deposition in
25 eastern Canada. Lovett and Lindberg (1986) concluded that dry deposition of nitrate is the
26 largest form of inorganic nitrogen deposited to oak-hickory forests in eastern Tennessee.
27 Annual estimates of NH3 deposition have been reported (Cowling and Lockyer 1981; Sinclair
28 and van Houtte, 1982), but numerous reports of NH3 evolution from foliage under conditions
29 of high soil N confound simple estimates of annual NH3-N deposition. Lovett (1991)
30 summarized research data for a number of forested sites in North America and Norway and
August 1991
10-209 ' DRAFT-DO NOT QUOTE OR CITE
-------
1 ' concluded that dry deposition of nitrogen typically occurs at annual rates approximately equal
2 to nitrogen deposited in precipitation.
3 Because gaseous deposition is difficult to measure accurately or continuously at the
4 landscape level of resolution, estimates of dry nitrogen deposition must rely on models.
5 Rigorous models of pollutant deposition have been developed (Hicks et al., 1985; Baldocchi,
6 1988; Baldocchi et al., 1987), and will be needed in the future for accurate determination of
7 reactive nitrogen gas and particle deposition to forest stands and ecosystems. Although
8 progress has been made in understanding the modeling process that control the dry deposition
9 of nitrogen containing compounds, additional research will be required to minimize errors in
10 predictions of total dry nitrogen deposition to specific regions and under a range of
11 environmental conditions.
12 Increased efforts have been made to establish both wet and dry deposition rates of
13 nitrogen to various types of ecosystems. These current deposition data are important as they
14 provide a basis to evaluate potential effects against "suggested critical levels". Although the
15 concept of critical nitrogen loading has not yet been widely adopted in North America (based
16 on amount of published data), a comparison of total nitrogen deposition data for North
17 America and proposed critical loads provide a reasonable comparison of the status of
18 terrestrial systems with respect to changes expected from elevated levels of nitrogen
19 deposition. Table 10-31 summarizes wet deposition data for nitrate and ammonium in the
20 United States. Since the data are for wet deposited forms of nitrogen, they represent an
21 underestimate of the total nitrogen deposition to the ecoystems. Table 10-32 summarizes
22 information regarding the total (wet and dry) deposition of nitrogen to a variety of
23 ecosystems/forest types or regional areas in North America and Europe.
24
25 10.7.4 Effect of Deposited Nitrogen on Forest Vegetation and Soils
26 The effects of N deposition upon biological systems must be viewed from the
27 perspective of the amount of N in the system, the biological demand for N, and the amount
28 of deposition. If N is deposited on an N-deficient ecosystem, a growth increase will likely
29 occur. If N is deposited on an ecosystem with adequate supplies of N, nitrate leaching will
30 eventually occur. Nitrate leaching is usually deemed undesireable in that it can contaminate
31 groundwater and lead to soil acidification.
August 1991
10-210
DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-31. MEAN ANNUAL WET NITRATE AND AMMONIUM DEPOSITION
TO VARIOUS STATES LOCATED THROUGHOUT THE UNITED STATES.
THE STATES ARE PRESENTED IN ORDER OF THE GREATEST
ANNUAL N DEPOSITION
Location
Pennsylvania
New York
Ohio
Georgia
Tennessee
Illinois
N. Carolina
Arkansas
Virginia
Florida
Oklahoma
Colorado
Alabama
New Mexico
S. Dakota
Texas
California
Washington
Wyoming
Arizona
Utah
Idaho
Oregon
Montana
Arizona
Hawaii
No. of Sites
3
5
2
1
1
4
4
1
1
2
3
4
1
4
1
3
.5
3
3
1
1
1 '
4
4
1
1
Forms
Nitrate
10.9
9.7
7.6
.6.9
6.9
6.2
6.2
5.0
5.3
4.9
4.1
4.3
3.7
3.6
2.7
3.1
2.9
2.7
2.5
2.6
2.5 '
2.3
2.1
1.9
1.0
0.08
of N Deposition
Ammonium
1.3
'"' 1.4 " '
'1.7
1.1
0.8
1.3
1.1
1.3
0.5
0.6
1.3
0.6
0.6
0.5
1.3
0.6
0.6
0.3
0.4
0.2
0.3
0.3
0.3
,0.4
0.2
0.01
(Kg ha'1 yr'1)
Totala
12.2
11.1
9.3
8.0
7.7
7.5
7.3
6.3
5.8
5.5
5.4
4.9
4.3
4.1
4.0
3.7
3.5
3.0
2.9
2.8
2.8
2.6
2.4
2.3
1.2
0.1
Total deposition data is for wet deposited forms only and as such represents an underestimate of the total
nitrogen loading received by these geographic areas.
August 1991
10-211 DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-32. MEASUREMENTS OF VARIOUS FORMS OF ANNUAL NITROGEN
DEPOSITION TO NORTH AMERICAN AND EUROPEAN ECOSYSTEMS.
MEASUREMENTS OF TOTAL DEPOSITION DATA THAT DO NOT INCLUDE
BOTH A WET AND DRY ESTIMATE PROBABLY UNDERESTIMATE TOTAL
NITROGEN DEPOSITION AND ARE ENCLOSED IN PARENTHESES
Site
Location/Vegetation
Forms of N Deposition (Kg ha"1 yr"1)
Wet . . . Dry
Cloud Rain
Particles Gases
Total
United States
CA, Chaparral
CA, Sierra Nevada
GA, Loblolly pine
NC, Hardwoods
NC, White pine
NY, Red spruce
TN, Oak forest #2
WA, Douglas fir
U.S. Regions
Adirondacks
Canada
Alberta (southern)
British Columbia
Ontario
Ontario (southern)
Fed. Rep. Germany
Spruce (SE slope)
Netherlands
Oak-birch
Douglas fir
Douglas fir
Norway
Spruce
8.2
—
3.7
4.8
3.7
7.3 6.1
6.0
1
6.3
7.3
5.5 '
3.7
2.3
16.5
_.
._
19.3
10.3
-
__
1.0 4.2
0.5
0.9 2.7
0.2 2.3
1.2
—
4.7
12.2C
—
-
1.4
__
_.
—
95.7°
0.7 0.2
23b
•(2)
9
5.3
7
16
7
(1)
'
11
19.5
(5)
(4)
3.7
16.5
24-56b
17-64b
115
11.2
'—Symbolizes data not available or in the case of cloud deposition not present.
'Total nitrogen deposition was based on bulk deposition and througbfall measurements and does include
components of wet and dry deposition.
Includes deposition from gaseous forms.
August 1991
10-212 DRAFT-DO NOT QUOTE OR CITE
-------
1 This analysis focuses upon forest ecosystems, because they are sensitive to both the
2 positive and negative effects of N deposition. Agricultural lands are excluded from this
3 discussion because crops are routinely fertilized with amounts of N (100-300 kg/ha) that far
4 exceed pollutant inputs even in the most heavily polluted areas. Pollutant N inputs to
5 grasslands and arid soils can be expected to produce increased growth in some instances,
6 . despite water limitations (e.g., Fisher et al., 1988). However, these systems are obviously
7 not subject to the soil acidification and groundwater NO3" pollution problems that might
8 occur in more humid areas. Excess N deposited on these ecosystems leaves via either
9 denitrification or NH4+ volatilization (see review by Woodmansee, 1978).
10 The biological competition for atmospherically-deposited N among heterotrophs
11 (decomposing microorganisms), plants, and nitrifying bacteria combined with the chemical
12 reactions between NH4+ and humus in the soil determine the degree to which vegetation
13 growth increase will occur and the degree to which incoming N is retained within the
14 ecosystem. Until recently, nitrifying bacteria were thought to be poor competitors for N,
15 with heterotrophs the most effective competitors and plants intermediate. Recent studies of
16 soil N dynamics using 15N (Davidson et al., 1990) and a through analyses of forest N
17 budgets suggest that these assumptions and perhaps our conceptual model of soil N Cycling
18 need modification. Specifically, nitrification may be proceeding at a significant level without
19 the appearance of NO3~ in soils or soil solution if NO3" is rapidly taken up by heterotrophs.
20 It is also clear that trees can be very effective competitors for atmospherically-deposited N in
21 N-deficient ecosystems. Finally, the role of chemical reactions between NH4+ and humus
22 need to be investigated; such reactions have been shown to be very important in fertilization
23 studies, and they may also play a major role in unfertilized ecosystems. If this is the case,
24 the fundamental assumption that N retention is controlled primarily by biological processes
25 may be erroneous.
26 Nitrification and NO3" leaching become significant only after heterotroph and plant
27 demand for N are substantially satisfied, a condition that has been referred to as "nitrogen-
28 saturated" or "N-saturated". Additions of N in any biologically-available form (NH4 ,
29 NO3", or organic) to an N-saturated system will cause equivalent leaching of NO3", except in
30 those very rare systems where nitrification is inhibited by factors other than competition from
31 heterotrophs and plants. Considering the effects of NO3' only will result in a substantial
August 1991
10-213 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
underestimation of the acidification potential of atmospheric deposition in N-saturated
ecosystems.
Vegetation demand for N depends upon a number of growth-influencing factors
including temperature, moisture, availibility of other nutrients, and stand age. Uptake rates
decline as forests mature, especially after the cessation of the buildup of nutrient-rich foliar
biomass following crown closure. Thus, N-saturation tends to be more common in older
forests than in younger forests. Processes that cause net N export from ecosystems such as
fire and harvesting will naturally push ecosystems toward a state of lower N-saturation or
even N deficiency. Intense fire causes a large loss of ecosystem N capital, but frequent, low
intensity fires may have little effect.
A review of the literature on forest fertilization and N cycling studies under various
levels of pollutant N input reveals some interesting contrasts that pertain to the the relative
roles of heterotrophs, plants, and nitrifiers discussed above. Forest fertilization has proven
quite successful in producing growth increases in N-deficient forests, even though trees
typically recover only 5-50% of fertilizer N (Table 10-33). On an ecosystem level, however,
retention of N is usually quite high (often 70-90% of applied N; Table 10-33), primarily due
to fertilizer N retention in the litter and soil, including non-biological reactions between
NH4* and humus. Fertilization studies differ from pollutant N deposition in several
important respects: (1) pollutant N deposition enters the ecosystem at the canopy level
whereas fertilizer is typically (but not always) applied to the soil, (2) fertilization leads to
high concentrations of NH4+ and, in the case of urea, high pH, both of which are conducive
to non-biological reactions between soil humus and NH4+, and (3) pollutant N deposition
enters the ecosystem as a slow, steady input in rather low concentrations, whereas the
fertilizer is typically applied in 1-5 large doses. Both plants and nitrifying bacteria are
favored by slow, steady inputs of N, possibly giving them a competative advantage over
heterotrophs for pollutant N inputs. A review of the literature on N cycling in unfertilized
forests with differing levels of pollutant N input supports this hypothesis. Ecosystem-level
recovery of atmospherically-deposited N (typically less than 50% and often 0%; Table 10-34
and Figure 10-33) is lower than of fertilizer N (typically 70-90% of applied N; Table 10-33
and Figure 10-34). It also appears as if vegetation retention of incoming N in unfertilized
forests is somewhat higher than in fertilized forests whereas soil (heterotroph) retention of
August 1991
10-214 DRAFT-DO NOT QUOTE OR CITE
-------
I
PQ
O
o
CO
c
r-t
S
w
i ^'1'
3
# °
S «
jd g J'
S3 g 6b s
/»> K> M«
CO
oo oo oo oo oo oo
ON ON ON ON ON ON
oo ON
en en en en en
VO vo vo vo vo
~ - ON ON ON
S3 S3 S3 S- v-' 73 T5 T5
T™^ *""! ^3 r^!i *rt W3 K3 W
•g s s 8
fS O O O
T3. ^ irH ta "3 "S
§ §
8 g
U2 W
33 03
O O
•rt -rt
\T>.
O >O vo ON ON t>
\o •* t> t>
O oo oo en
eN (N \o en
O en en O CM
vo
rf O
CN •*
m O
CM v>
00 O)
O i-<
o »/•>
000
IT) O O
^ en vo
O
v>
80
•* •*
CM CM
ca cs
O O O
O O O
en CN f»
O O
eSl CN
O O
cS o en en en en
&^ c^
-------
August 1991
10-216 DRAFT-DO NOT QUOTE OR CITE
-------
Ecosystem N Retention vs Fertilizer N Input
2000
to
c
O
1000 -
y = 58.682 + 0.51747x R*2 - 0.592
1000
2000
Fertilizer Input (kg/ha)
Figure 10-33. Ecosystem recovery of fertilizer N as a function of fertilizer N input.
1 atmospherically-deposited N is much lower. In forests with very low atmospheric N inputs,
2 it appears as if the soil is being "mined" for the N necessary to supply vegetation, an
3 indication that plants are actually out-competing heterotrophs for N. In forests with high
4 atmospheric N inputs, heterotrophic N uptake appears to be minimal, perhaps because of
5 limitations by organic substrates or other nutrients.
August 1991
10-217 DRAFT-DO NOT QUOTE OR CITE
-------
IX
"m
^
a
jo
c
o
o
tr
Ecosystem N Retention vs Atmospheric N Input
30
20-
10-
•10-
-20-
-30'
•O
—T—
20
—r-
40
80
Atmospheric N Input (kg/ha/yr)
Figure 10-34. Ecosystem N retention as a function of atmospheric N input.
1
2
3
4
5
6
Since nitrification results in the creation of nitric acid within the soil, there are concerns
that elevated N inputs to N-saturated systems will result in soil acidification and aluminum
mobilization. There are very few proven, documented cases in which excessive atmospheric
N deposition has caused soil acidification (e.g., in forests in The Netherlands subject to very
high N deposition levels), but there is no doubt that the potential exists for many mature
forests with low uptake rates, given high enough inputs for a sufficiently long time. The
August 1991
10-218
DRAFT-DO NOT QUOTE OR CITE
-------
1 greatest uncertainty in assessing and projecting rates of soil acidification is the estimation of
2 weathering rates (i.e., the release of base cations from primary minerals).
3 Soil acidification is usually thought of as an undesireable effect, but in some cases, the
4 benefits of alleviating N deficiency clearly outweigh the detriments of soil acidification
5 (e.g., the benefits of N-fixation by red alder always outweigh the detriments of soil
6 acidification to succeeding Douglas fir stands in the Pacific Northwest).
7 Increased concentrations of NO3" or any other mineral acid anion (e.g., SO42", or Cl")
8 in soil solution lead to increases in the concentrations of all cations in order to maintain
9 charge balance in solution. Equations describing cation exchange in soils dictate that as the
10 total anion (and cation) concentrations increase, individual cation concentrations increase as
11 follows: A13+ > Ca2+, Mg2+ > K+, Na+, H+. Thus, soil solution A13+ concentrations
12 increase not only as the soil acidifies (i.e., as the proportion of A13+ on the exchange
13 complex increases) but also as the total ionic concentration of soil solution increases.
14 There are several cases in which A13+ concentrations in natural waters have been shown
"2 _i_
15 to be positively correlated with NO3" concentrations. Ulrich (1983) noted NO3" - Al
16 pulses in soil solutions from the Soiling site in Germany during warm dry years. He
17 hypothesized that these nitrate-induced A13+ pulses caused root damage and and were a major
18 contributor to forest decline observed in Germany during the mid 1980's. This hypothesis is
19 disputed by other German forest scientists who point out that forest decline occurred on base-
20 rich as well as base-poor soils (the base rich soils not being subject to A13+ pulses) (e.g.,
21 Rehfuess, 1987). Van Breemen et al. (1982, 1987) and Johnson et al. (in press) note NO3" -
22 A13+ pulses in soil solutions from forest sites in The Netherlands and in the Smoky
23 Mountains of North Carolina. Aluminum toxicity is one of several nitrogen-related
24 hypotheses posed to explain forest decline in both countries. Other hypotheses include
25 weather extremes and climate change, Mg and K deficiencies which occur in sites naturally
26 low in these nutrients, and foliar damage due to acid mist. Researchers on aquatic effects of
27 • acid deposition have long noted springtime pulses of NO3', A13 + , and H+ in acid-affected
28 surface waters of the Northeastern U.S. (Galloway et al., 1980; Driscoll et al., 1989).
29 Increased N deposition can cause significant changes in tree physiological function,
30 susceptibility to insect and disease attack, and even plant community structure. Several
31 hypotheses posed to explain current forest declines in eastern North America invoke the
August 1991
10-219 DRAFT-DO NOT QUOTE OR CITE
-------
1 effects of excess N deposition upon physiological processes. These physiological responses
2 generally invoke altered carbohydrate allocation causing increased sensitivity to drought,
3 frost, or insect attack. To date, however, experimental evidence has not supported these
4 hypotheses.
5 Excessive NH4+ deposition to soils in which nitrification is inhibited causes serious
6 nutrional imbalances and even toxic effects to some forests in The Netherlands (Boxman
7 et al., 1988). Deleterious effects of excess N deposition can occur via aboveground
8 processes as well: K and Mg deficiencies in declining Dutch forests are thought to be caused
9 by excessive foliar leaching due to high inputs of NH4+ (Roelofs et al., 1985).
10 Growth responses to increased N inputs may result in changes in species composition.
11 Species respond differentially to increased N availability, creating the potential for changes in
12 ecosystem composition with increased N loading. Changes from heathland to grassland in
13 Holland have been attributed to current rates of N deposition (Roelofs et al., 1987).
14 Ellenberg (1987) points to further species changes in Central European ecosystems as a likely
15 consequence of elevated N. He states that "More than 50% of the plant species in Central
16 Europe can only compete on stands that are deficient in nitrogen supply".
17 Increased N inputs can affect tree resistance to insect and desease either positively or
18 negatively. Alleviating N deficiency may increase plant resistance to pathogen attack, but it
19 may also reduce the production of phenols in plant tissues, thereby reducing resistance to
20 pathogen attack. To date, there is little research to show how increased N inputs affect
21 susceptibility to pathogen attack, but the potential for either positive or negative effects is
22 significant.
23
24 10.7.5 Effects of Nitrogen on Terrestrial Vegetation
25 Interpretation of the effects of wet and dry deposited nitrogen compounds at the
26 ecosystem level is difficult because of the interconversion of nitrogen compounds and the
2,7 complex interactions which exist between biological, physicochemical, and climatic factors.
28 Nevertheless, reactive nitrogen compounds have been hypothesized to impact ecosystems
29 directly through modifications of individual plant physiological processes, or indirectly
30 through alterations in the nitrogen status of the ecosystem.
August 1991
10-220
DRAFT-DO NOT QUOTE OR CITE
-------
1 Very little information is available on the direct effects of nitric acid vapor on
2 vegetation and essentially no information on its effects on ecosystems. Norby et al. (1989)
3 reported that nitric acid vapor (0.075 ppmv) induced nitrate reductase activity (NRA) in red
4 spruce foliage. The effects of ammonia, a reduced nitrogen gas, have been summarized by
5 Van der Eerden (1982), however, ammonia concentrations seldom reach phytotoxic levels in
6 the United States (U.S. Environmental Protection Agency 1982). In contrast, high ammonia
7 concentrations in Europe have been observed (van Dijk and Roelofs, 1988). Van der Eerden
8 (1982) summarized available information on the direct response of crop and tree species to
9 ammonia fumigation and concluded that the following concentrations produced no adverse
10 effects:
11 0.107 ppmv (75/ig m'3), yearly average
12 0.858 ppmv (600 Mg m~3), daily average
13 14.3 ppmv (10,000 ng m'3) hourly average.
14
15 Submicron, ammonium sulfate aerosols have been shown to affect foliage of Phaseolus
•3
16 vulgaris L. (Gmur et al., 1983). Three-week exposure to a concentration of 26 mg m"
17 (37 ppmv) produced leaf chlorosis, necrosis and loss of turgor.
18 Because current ambient concentrations of NO, NO2 and NH3 are low across much of
19 the United States except in certain highly populated urban areas, significant direct effects of
20 these nitrogen compounds on ecosystems seems unlikely at the current time. Concentration
21 and effects data are unavailable for making similar conclusions regarding other reactive
22 nitrogen compounds like nitric acid vapor or the gaseous nitrate radical.
23 Serious consideration is currently being given to hypotheses that excess total nitrogen
24 deposition may impact plant productivity directly or through changes in soil chemical
25 properties. Furthermore it has been proposed that excess nitrogen deposition to ecosystems
26 may be modifying interplant competitive balances leading to changes in species composition
27 and/or diversity.
28 De Temmerman et al. (1988) found increased fungal outbreaks and frost damage on
29 several pines species exposed to very high ammonia deposition rates (>350 kg ha" yr" ).
30 Numbers of species and fruiting bodies of fungi have also increased concomitantly with
31 nitrogen deposition in Dutch forests (van Breemen and van Dijk, 1988). Schulze (1989)
32 presents a clear progression of evidence which indicates that canopy uptake of nitrogen
August 1991
10-221
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
together with root uptake has caused a nitrogen imbalance in Norway spruce leading to forest
decline.
Excessive nitrogen inputs to terrestrial ecosystems can cause differential competitive
advantage among plants within a heathland (Heil and Bruggink, 1987; Heil et al., 1988). In
unmanaged heathlands in The Netherlands, Calluna vulagris is being replaced by grass
species, as a consequnce of the eutrophic effect of acidic rainfall and large nitrogen inputs
arising from intensive farming practices in the region. Calluna is an evergreen with a long
growing season which normally permits it to compensate for its slow growth rate, so that it
competes successfully with the faster growing Molinia (grass) under normal nutrient-limiting
conditions. However, a large increase in the nitrogen supply improves the competitive
advantage of Molinia, increasing its growth rate so that it becomes the dominant species in
the heathland. Roelofs et al. (1987) observed that nitrophilous grasses (Molinia and
Deschampsid) are displacing slower growing plants (Erica and Calluna) on heathlands in The
Netherlands, and suggested that a correlation existed between this change and nitrogen
loading. Van Breemen and van Dijk (1988) found a substantial displacement of heathland ,
plants by grasses from 1980 to 1986 and also observed increases in nitrophilous plants in
forest herb layers. Ellenberg (1988) suggested that ionic inputs (NO3~ and NH4+) influence
competition between organisms long before toxic effects appear on individual plants. These
changes in The Netherlands have occurred under nitrogen loadings of between 20 and 60 kg
N ha"1 yr"1. Liljelund and Torstensson (1988) have shown clear signs of vegetation changes
in response to nitrogen deposition rates of 20 kg ha"1 yr"1.
10.7.6 N Saturation, Critical Loads, and Current Deposition
Ecosystem nitrogen saturation and the definition of the critical N deposition levels have
been the subject of recent conferences (Nilsson and Grennfelt,. 1988; Brown et al., 1988;
Skeffington and Wilson, 1988). Miller and Miller (1988) proposed three definitions for
N saturated ecosystems:
1. No response to additional N,
2. Growth reductions in response to added N or,
3. Added N leads to'increased losses of nitrate in streamwater.
August 1991
10-222 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7 '<
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22.
23
24
25
26
27
28
29
30
31
32
33
34
They concluded that third was most reasonable'. Brown et al. (1988), however, reported
that a recent workshop, based on a model of plant/soil N saturation put forth by Agren and
Bosatta (1988), concluded that N saturation could best be defined as occurring when N losses
from ecosystems exceeded inputs. Aber et al. (1989) similarly define nitrogen saturation as
the availability of ammonium and nitrate in excess of total combined- plant and microbial
nutritional demands. The concept of nitrogen saturation makes it possible to define a critical
N load (deposition rate) at which no change or deleterious impacts will occur to an ecosystem
(Nilsson, 1986). It is important to recognize that the magnitude of such a "critical load" will
be site and species specific because it is highly dependent on initial soil chemistries and
biological growth potentials (i.e., nitrogen demands). Skeffington and Wilson (1988) point
out that intrinsic in all definitions of a "critical load" is the notion that there is a load at
which no long-term effects occur. The complexity of the N cycle and ecosystem diversity
make defining a critical load for N very difficult. The following possible criteria may be
useful for defining appropriate critical N loads on ecosystems: .
1. ' Prevent nitrate levels in drinking or surface waters from rising above
standard levels
Ensure proton production less than weathering rate
Maintenance of a fixed ammonia-base cation balance
Maintenance N inputs below N outputs (the N saturation approach)
Minimize accelerations in the rates of ecological succession (vegetation changes
due to altered interspecific competition). ' !
2.
3.
4.
5.
De Vries (1988) defined criteria for a combined critical load for nitrogen and sulfur for
Dutch forest ecosystems using the following: N contents of foliage, nitrate concentrations in
groundwater, NH4/K ratios, Ca/Al ratios, and Al concentrations in soil solution. Based on
these criteria, De Vries concluded that current rates of N and S deposition in The Netherlands
exceed acceptable levels.
Schulze et al. (1989) proposed critical loads for N deposition based on an ecosystem
total anion and cation balance. This approach makes the assumption that processes
determining ecosystem stability are related to soil acidification and nitrate leaching (see
Section 10.3.6). They concluded that in order to limit the mobilization of aluminum and
other heavy metals resulting from acidification and nitrate leaching (a negative result), critical
nitrogen deposition rates could not exceed 3-14 Kg N ha"1 yr"1 for silicate soils or 3 to
48 kg N ha"1 yr"1 for calcareous based soils. Other critical loads have been proposed at rates
August 1991
10-223 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
of nitrogen deposition ranging from as little as 1 to levels near 100 Kg N ha'1 yr"1 depending
on the impacts considered acceptable and the criteria used to define the critical load.
Using criteria to minimize species change, critical loads less than 30 kg ha"1 yr"1 have
been proposed (van Breeman and van Dijk, 1988; Liljelund and Torstensson, 1988). When
using the criteria that ecosystem nitrogen inputs should not exceed outputs, critical loads have
been proposed as low as 1-5 kg N ha"1 yr"1 for poor productive sites with low productivity or
in the range from 5-30 kg N ha"1 yr"1 for sites having medium quality soils and for common
forested systems (Boxman et al., 1988; Rosen, 1988; Skeffington and Wilson, 1988; World
Health Organization, 1987).
In summarizing the results of a recent conference on critical nitrogen loading, after
discussing various options for setting a critical N load, Skeffington and Wilson (1988)
concluded that "we do not understand ecosystems well enough to set a critical load for
N deposition in a completely objective fashion." Brown et al. (1988) further concluded that
there was probably no universal critical load definition that could be applied to all
ecosystems, and a combination of scientific, political, and economic considerations would be
required for the application of the critical load concept.
The following terrestrial ecosystems have been suggested as being at risk from the
deposition of nitrogen-based compounds:
• heathlands with a high proportion of lichen cover,
• low meadow vegetation types used for extensive grazing and
haymaking, and
• coniferous forests, especially those at high altitudes (World Health
Organization, 1987).
The above oligotrophic ecosystems are considered at risk from atmospheric nitrogen
deposition because plant species normally restricted by low nutrient concentrations could gain
a competitive advantage, and their growth at the expense of existing species would change the
"normal" species composition and displace some species entirely (Ellenberg, 1987; Waring,
1987). Sensitive natural ecosystems, unlike highly manipulated agricultural systems, may be
prone to change from exposure to dry deposited nitrogen compounds because processes of
natural selection whereby tolerant individuals survive may not be keeping pace with the
current levels of atmospheric N deposition (World Health Organization, 1987).
August 1991
10-224 DRAFT-DO NOT QUOTE OR CITE
-------
1 There is little doubt that N deposition has a prounced effect on many if not most
2 terrestrial ecosystems. Since most forest ecosystems in North America are N deficient, one
3 of the most noticeable initial changes in response to increased N deposition is likely to be a
4 growth increase. Whether this growth increase is deemed desireable or undesireable in a
5 particular ecosystem is entirely a matter of management objectives (timber production or
6 wilderness preservation), and, ultimately, value judgements by, society, For instance,
7 improved growth and vitality due to increased N deposition may not be deemed desireable in
8 wilderness areas. A blanket statement as to the benefit or damage due to N deposition cannot
9 be made, but it is logical to assume that increased N deposition would be considered
10 beneficial, on balance, to most North American forests. Exceptions to this will certainly
11 occur, and have already been noted in specific situations such as high-elevation forests with
12 low vegetation demand and high atmospheric N input.
13 However, because ecosystems have a variable capacity to buffer changes caused by
14 elevated inputs of nitrogen, it is difficult to make general conclusions about the type and
15 extent of change (if any) currently,resulting from nitrogen deposition in North America.
16 More research needs to be conducted in this area to determine if the hypothesized effects of
17 excess nitrogen deposition are taking place and to determine the sensitivity of a wide range of
18 natural ecosystems to nitrogen loading.
19
20 10.7.7 Effects of Nitrogen on Wetlands and Bogs
21 The anaerobic (oxygen-free) nature of their waterlogged soils is the feature that sets
22 wetlands apart. Anaerobic wetland soils favor the accumulation of organic matter and losses
23 of mineral nitrogen to the atmosphere through denitrification reactions (the conversion of
24 nitrate to gaseous nitrogen by microbes). Nitrogen deposition can impact plant and microbial
25 processes either directly or indirectly by acidifying the environment. An increase in nitrogen
26 supply through atmospheric deposition or other means alters the competitive relationships
27 among plant species such that fast growing nitrophilous species (species that have a high
28 nitrogen requirement) are favored. Microbial rates of decomposition, nitrogen fixation (the
29 conversion of gaseous nitrogen to ammonium), nitrification (the conversion of ammonium to
30 nitrate), and dissimilatory nitrate reduction (conversion to gaseous nitrogen or ammonium)
31 are all affected. Acidification below pH 4 to 5.7 blocks the nitrogen cycle by inhibiting
August 1991
10-225
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
nitrification, and the accumulation of NH4+ (ammonium) in the environment represses
nitrogen fixation (Roelofs, 1986; Schuurkes et al., 1986, 1987; Rudd et al., 1988). The
proportion of N2O (nitrous oxide) produced by denitrification reactions increases with
decreasing pH below 7, and the absolute rate of production of N2O increases with increasing
eutrophication (nutrient enrichment of the environment) (Focht, 1974). This is potentially
important on a global scale because of chemical reactions with N2O in the atmosphere that
result in a loss of ozone. • , .
The importance of atmospheric nitrogen deposition to the community structure (species
composition and interrelationships) of wetlands increases as rainfall increases as a fraction of
the total water budget. Primary production (plant growth) in wetlands is commonly limited
by nitrogen availability. Primary production is proportional to the rate of internal nitrogen
cycling, which is influenced by the quantity of mineralizable soil nitrogen as well as the
supply of nitrogen to the ecosystem from the atmosphere or surface flow. Total nitrogen
inputs range from about 10 kg N ha"1 yr"1 in ombrotrophic bogs (rain-fed bogs), which
receive water only through precipitation, to 750 kg N ha"1 yr"1 or more in intertidal wetlands
with large ground and surface hydrologic inputs.
From studies of nine North American wetlands, bulk nitrogen deposition ranges from
5.5 to 12 kg N ha"1 yr"1 and occurs in the form of NO3" (nitrate), NH4+ (ammonium), and
dissolved organic nitrogen in roughly equal proportions. More recent studies, however,
suggest that these rates are too low and that the wet deposition of NQ3" alone is greater than
15 kg N ha"1 yr"1 over much of eastern North America (Zemba et al., 1988), Dry
deposition, which probably accounts for greater than 50% of total deposition, adds to the
total. Leaf-capture of nitrogen in fog droplets is a third form of deposition that is locally
important. Applications of nitrogen fertilizer in the field, ranging from 7 to
3,120 kg N ha"1 yr"1, have increased standing biomass by 6 to 413%. Other nutrients, like
phosphorus, become secondarily limiting to primary production after nitrogen inputs reach a
threshold. Fertilization and increased atmospheric deposition have increased the dominance
of grass species over other plant species in bogs,, and extreme eutrophication is associated
with a decrease in plant species diversity.
Single additions to vegetated wetland soils of 15N-labelled mineral nitrogen at rates of
about 100 kg N ha"1 yr"1 indicate that up to 93% of applied NH4+ is rapidly assimilated into
August 1991
10-226
DRAFT-DO NOT QUOTE OR CITE
-------
1 organic matter within a single growing season. The majority of the labelled nitrogen is lost
2 from the system after 3 years by the combined processes of advective transport in water
3 (carried in moving water) of paniculate organic matter, advective and' diffusive transport of
4 dissolved nitrogen, and denitrification. In the absence of plants, the major fate of inorganic
5 nitrogen applied to wetland soils is loss to the atmosphere by denitrification.
6 Peat-forming Sphagnum spp. are largely absent from bogs in western Europe where
7 bulk deposition rates are about 20-40 kg N ha"1 yr"1, and soft water communities once
8 dominated by isoetids in The Netherlands have been converted to later successional stages
9 dominated by Juncus spp. (rush) and Sphagnum spp. or to grasslands. Heathlands dominated
10 by shrubs have also converted to grasslands. Experimental studies indicate mat ombrotrophic
11 bogs can be maintained if nitrogen inputs are less than 20 kg N ha"4 yr"1. Increased
12 productivity associated with eutrophication is accompanied by increased rates of transpiration
13 (evaporation of water from leaf surfaces) which can alter wetland hydrology and influence the
14 direction of wetland succession. By this mechanism, one modelling study suggests that a
15 succession (change) from open ombrotrophic bog to forested wetland occurs when a threshold
16 of 7 kg N ha"1 yr"1 is exceeded. These estimates are consistent with conclusions from studies
17 of species distributions that place the limit for many species from 10 to 20 kg N ha"1 yr"1
18 (Liljelund and Torstensson, 1988).
19 . Fourteen percent (18 species) of the plant species from the conterminous United States
20 ' that are formally listed as endangered, and an additional 284 species listed as potentially
21 threatened (Code of Federal Regulations, 1987), are found principally in wetland habitats.
22 Some of the endangered plants, like the green pitcher plant, are known to be adapted to
23 infertile habitats and are threatened by current levels of nitrogen deposition in parts of North
24 America. Plant species that are threatened by high nitrogen deposition are not confined to
25 wetland habitats, however, but are common across many ecosystem types (Ellenberg, 1988).
26 ' '• - ' '
27 10.7.8 Effects of Nitrogen on Aquatic Systems
28; Nitrogen deposition has not historically been considered a serious threat to the integrity
29 of aquatic ecosystems.
30 Assessment of the aquatic effects of nitrogen oxides depends on a close examination of
31 'the processes by which nitrogen may enter streams, lakes and estuaries. Sources of nitrogen
August 1991
10-227
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
may include: (1) atmospheric deposition directly to the water surface; (2) deposition to the
watershed that is subsequently routed to the drainage waters; (3) gaseous uptake by plants that
is subsequently routed, by way of litter fall and decomposition, to drainage waters; and
(4) nitrogen fixation, either in the water itself, or in watershed soils. In addition, numerous
processes act to transform nitrogen species into forms that are only indirectly related to the
original deposition or fixation. These transformations include: (1) nitrogen assimilation (the
biological uptake of inorganic nitrogen species); (2) nitrification (the oxidation of ammonium
to nitrate); (3) denitrification (the biological reduction of nitrate to form gaseous forms of
nitrogen, N2, NO, or N2O); and (4) mineralization (the decomposition of organic forms of
nitrogen to form ammonium). The multiple sources of nitrogen to aquatic systems, and the
complexities of nitrogen transformations in water and watersheds, have the effect of
de-coupling nitrogen deposition from nitrogen effects, and reduce our ability to attribute
known aquatic effects to known rates of nitrogen deposition. While it is not currently
possible to trace the pathway of nitrogen from deposition through any given watershed and
into drainage waters, we can, in areas of the United States where non-atmospheric sources of
nitrogen are small, begin to infer cases where nitrogen deposition is having an impact on
aquatic ecosystems.
Any discussion of the aquatic effects of nitrogen oxides must focus on the concept of
nitrogen saturation. Nitrogen saturation can be defined as a situation where the supply of
nitrogenous compounds from the atmosphere exceeds the demand for these compounds on the
part of watershed plants and microbes (Skeffmgton and Wilson, 1988; Aber et al., 1989).
Under conditions of nitrogen saturation, forested watersheds that previously retained nearly
all of nitrogen inputs, due to a high demand for nitrogen by plants and microbes, begin to
supply more nitrogen to the surface waters that drain them. Our conceptual understanding of
nitrogen saturation suggests that, in aquatic systems, the earliest stages of nitrogen saturation
will be observable as increases ;in the severity and duration of springtime pulses of nitrate.
The aquatic effects of nitrogen oxides can be divided into three general categories:
(1) acidification, both chronic and episodic; (2) eutrophication of both freshwaters and
estuaries; and (3) directly toxic effects.
August 1991
10-228 DRAFT-DO NOT QUOTE OR CITE
-------
1 10.7.8.1 Acidification
2 Acidification effects are traditionally divided into chronic (long-term) and episodic
3 (short-term effects usually observable only during seasons of high runoff) effects. Nitrate,
4 the dominant form of inorganic nitrogen in almost all aquatic systems, is commonly present
5 in measurable concentrations only during winter and early spring, when terrestrial demand for
6 nitrogen is low because plants in the watershed are dormant. Nitrogen will therefore only be
7 a problem in chronic acidification in rare cases where the process of nitrogen saturation is
8 very much progressed. Chronic acidification by nitrogen can be conclusively demonstrated
9 only in parts of Europe (e.g., Hauhs, 1989; Hauhs et al., 1989; van Breemen and van Dijk,
10 1988).
11 . Episodic acidification by nitrate is far more common than chronic acidification, and is
12 well-documented for streams (Driscoll et al., 1987b) and lakes (Galloway et al., 1980;
13 Driscoll et al., 1991; Schaefer et al., 1990) in the Adirondack Mountains, for streams in the
14 Catskill Mountains (Stoddard and Murdoch, 1991; Murdoch and Stoddard, in review), and in
15 a small proportion of lakes in Vermont (Stoddard and Kellogg, in press), as well as in many
16 parts of Canada (Jeffries, 1990) and Europe (e.g., Hauhs et al., 1989).
17 Based on intensive monitoring data, it is possible to divide lakes and streams into three
18 groups, based on their seasonal NO3" behavior. In many parts of the country, nitrogen
19 demand on the part of the terrestrial ecosystem is sufficiently high that no leakage of NO3"
20 ' from watersheds occurs, even when nitrogen deposition rates are relatively high, and cold
21 temperatures should limit the biological demand for nitrogen. Lakes and streams in these
22 areas show no evidence that nitrogen deposition is producing adverse aquatic effects.
23 In a second group of lakes and streams, NO3~ concentrations show strong seasonally,
24 with peak concentrations during snow melt, or following large rain events. In many cases,
25 these episodic increases in NO3~, along with already low baseline acid neutralizing capacity
26 (ANC), are sufficient to cause short-term acidification and potential adverse biological
27 effects. It is important to note that seasonal increases in NO3~ concentrations can be
28 produced by normal watershed processes; lowered terrestrial demand for nitrogen during the
29 dormant season, for example, creates a strong likelihood that springtime drainage waters will
30 show NO3" concentrations that are elevated over summer and fall concentrations.
31 Mineralization of organic matter during the cold months of winter, coupled with low
August 1991
10-229
DRAFT-DO NOT QUOTE OR CITE
-------
1 biological demand for nitrogen, can produce high winter concentrations of NO3" in soil water
2 that is subsequently flushed into drainage waters during spring snow melt or during large rain
3 storms. While the seasonal pattern of elevated NO3" concentrations in this group of lakes and
4 streams can be considered normal, the severity of the NO3" episodes that these systems
5 experience can be strongly influenced by the amount of nitrogen stored in the snow pack over
6 the course of the winter. If biological demand for nitrogen is still low at the onset of snow
7 melt, the entire store of snowpack NO3" can be flushed into drainage waters in the very early
8 stages of snow melt (e.g., Johannessen and Henriksen, 1978; Jeffries, 1990).
9 The third group of lakes and streams exhibits both the strong seasonally in NO3"
10 concentration described in the previous paragraph, and increasing trends in NO3"
11 concentrations. Because the early stages of nitrogen saturation are expected to produce
12 increases in NO3" concentrations, especially during episodes, long-term increases in NO3"
13 may represent the strongest evidence that nitrogen deposition is responsible for aquatic
14 effects. In all cases where increasing trends in NO3" have been documented in the
15 United States (Smith et al., 1987; Stoddard and Murdoch, 1991; Murdoch and Stoddard, in
16 review; Driscoll et al., in review) they have occurred at a time when nitrogen deposition is
17 relatively constant (e.g., Simpson and Olsen, 1990). Increased leakage of NO3~ from
18 watersheds in these areas therefore represents a long-term decrease in the ability of
19 watersheds to retain nitrogen. A likely cause of such long-term changes is a lowering in the
20 demand for nitrogen as a nutrient on the part of the terrestrial ecosystem, which may result
21 from long-term high rates of nitrogen deposition to affected watersheds (e.g., Aber et al.,
22 1989), forest maturation (Elwood et al., 1991), or, more likely, a combination of both
23 factors.
24 The locations of lake and streams sites in each of the three NO3" groups are shown on
25 maps of the Northeast (Figure 10-35) the Southeast (Figure 10-36) and the West
26 (Figure 10-37). In order to assess which lake and stream sites fall into each group, it was
27 necessary to have data collected over several years (at least 3 years), and on a relatively
28 intensive sampling schedule (at least 4 times per year, to illustrate seasonal patterns). These *
29 criteria exclude many sources of data, most notable those from the National Surface Water
30 Survey (Linthurst et al., 1986; Landers et al., 1987; Kaufmann et al., 1988), and limit the
31 conclusions that can be drawn concerning the spatial extent of aquatic effects attributable to
August 1991
10-230
DRAFT-DO NOT QUOTE OR CITE
-------
o Data indicate no influence of NOj
• Data indicate strong influence of N
* Data indicate strong influence of NO"
and increasing trend in NO"
Pennsylvania
cP
Figure 10-35. Location of acid-sensitive lakes and streams in the northeastern United
States where the importance of NO3" to seasonal water chemistry can be
determined. Data from: Kahl et al. (1991); Wigington et al. (1989);
Driscoll et al. (1987a); DriscoU et al. (in review); Kramer et al. (1986);
Murdoch and Stoddard (in review); Eshleman and Hemond (1985);
Morgan and Good (1988); Baird et al. (1987); Likens (1985); Sharpe
et al. (1984); Stoddard and Kellogg (in press); DeWalle et al. (1988);
Barker and Witt (1990); Schofield et al. (1985); Phillips and Stewart
(1990).
1 nitrogen deposition. None-the-less, the maps illustrate the existence of severe problems in
2 the Northeast (especially the Adirondack and Catskill,Mountains), and the Southeast (in the
3 Mid-Appalachians and Great Smoky Mountains), and the potential for future problems in the
4 West.
August 1991
10-231 DRAFT-DO NOT QUOTE OR CITE
-------
O Data Indicate no Influence of NOJ
• Data Indicate strong influence of NO'
* Data Indicate strong influence of NO 3
and Increasing trend in NO'
N
Figure 10-36. Location of acid-sensitive lakes and streams in the southeastern United
States where the importance of NO3" to seasonal water chemistry can be
determined. Data from: Elwood et al. (1991); Cosby et al. (1991);
Elwood and Turner (1989); Buell and Peters (1988); Swank and Waide
(1988); Jones et al. (1983); Silsbee and Larson (1982); Katz et al. (1985);
Weller et al. (1986); Wigington et al. (1989); Kramer et al. (1986);
Edwards and Helvey (1991).
1
2
3
4
5
6
7
It is also possible to draw correlations between rates of nitrogen deposition, and rates of
»
nitrogen loss from watersheds; while these analyses cannot indicate causal relationships, they
can suggest patterns that merit further attention. Two independent attempts have been made
to relate deposition and watershed nitrogen export in the United States, and both suggest
similar conclusions. Kaufmann et al. (1991) used data from the National Stream Survey
(NSS; Kaufmann et al., 1988) and interpolated wet deposition values (of N03" + NH4+) to
correlate deposition and surface water dissolved inorganic nitrogen concentrations
August 1991
10-232 DRAFT-DO NOT QUOTE OR CITE
-------
o Data indicate no influence of NOj
• Data indicate strong influence of
* Data indicate strong influence of
and increasing trend in NOj
N
Figure 10-37. Location of acid-sensitive lakes and streams in the western United States
where the importance of NO3" to seasonal water chemistry can be
determined. Data from: Melack and Stoddard (1991); Stoddard (1987a);
Loranger et al. (1986); Wigmgton et al. (1989); Kramer et al. (1986);
Welch et al. (1986); Eilers et al. (1990); Gilbert et al. (1989).
1
2
3
4
5
(NO3~ + NH4+) in large physiographic regions of the eastern United States (Figure 10-38a).
The NSS was a probability-based sample of streams, sampled at spring base flow in 1987;
because it is probability-based, the results from the relatively small number of streams
sampled in the NSS can be extrapolated to the population of streams within each of the
9 regions sampled. The results of the correlation suggest a strong correspondence between
August 1991
10-233
DRAFT-DO NOT QUOTE OR CITE
-------
te
I
0-
100 200 300 400
Wet NCT + NH4* Deposition (eq/ha/yr)
500
f
tr
„
1
oc
if
i
I
350-
300-
250-
200-
150-
100-
50.
0-
o
0 0
°0
o
o
0
8 ° o
O o
O ° n ^ ° O
^-0 000? , °? ' ° °.
-------
1 median wet deposition of nitrogen in a region, and the median spring base-flow concentration
2 of nitrogen in a region. In addition, the results suggest a threshold rate of wet nitrogen
3 deposition of approximately 2.8 kg N ha"1 yr'1 above which significant losses of nitrogen
4 from watersheds can begin to occur.
5 Driscoll et al. (1989a) collected input/output budget data for a large number of
6 undisturbed forested watersheds in the United States and Canada, and summarized the
7 relationship between nitrogen export (of NO3") and wet nitrogen deposition (of
8 NO3" + NH4+). These data are supplemented in Figure 10-38b with some published
9 input/output data that were not included in the original figure. Driscoll et al. (1989a) stress
10 that the data were collected using widely differing methods and over various time scales
11 (from one year to several decades). Like the data of Kaufmann et al. (1991, Figure 10-38a),
12 these budget data suggest a threshold rate of wet nitrogen deposition of ca. 2.8 kg N ha'1 yr"1
13 above which significant export of NO3" from watersheds may occur.
14
15 10.7.8.2 Eutrophication
16 Assigning responsibility for the eutrophication of lakes and estuaries to nitrogen oxides
17 requires a determination of two key conditions. The first is that the productivity of the
18 aquatic system be limited by the availability of nitrogen, rather than by some other nutrient or
19 physical factor. The second is that nitrogen deposition be a significant source of nitrogen to
20 the system. In many cases of eutrophication, the supply of nitrogen from deposition is minor
21 when compared to other anthropogenic sources, such as pollution from either point or non-
22 point sources.
23 ; It is generally accepted that the productivity of fresh waters is limited by the availability
24 of phosphorus, rather than the availability of nitrogen (reviewed by Hecky and Kilham,
25 1988). Conditions of nitrogen limitation do occur in lakes, but are often either transitory, or
26 the result of high inputs of .phosphorus from anthropogenic sources. Often when nitrogen
27 - limitation does occur it is a short-term phenomenon, because nitrogen-deficient conditions
28 favor the growth of nitrogen-fixing blue green algae (e.g., Smith, 1982). Because nitrogen-
29 fixing species are not limited by the availability of fixed nitrogen (e.g., NH4+ or NO3") they
30 may thrive under conditions where other species are nitrogen limited, and effectively increase
31 rates of nitrogen input to the system (by fixation of gaseous nitrogen) beyond the levels
August 1991
10-235 DRAFT-DO NOT QUOTE OR CITE
-------
1 where system productivity can be said to be nitrogen limited. It appears that nitrogen
2 limitation may occur naturally (i.e., in the absence of anthropogenic phosphorus inputs) in
3 lakes with very low concentrations of both nitrogen and phosphorus, as are common in the
4 western United States, and in the Northeast. Suttle and Harrison (1988) and Stockner and
5 Shortreed (1988) suggest that phosphorus concentrations are too low in these systems to allow
6 blue green algae to thrive, because they are poor competitors for phosphorus at very low
7 concentrations. Results of the National Surface Water Survey (NSWS; Kanciruk et al., 1986;
8 Eilers et al., 1987) suggest that the largest number of potentially nitrogen limited lakes in the
9 United States occur in the West (20-30% of the population of lakes sampled by NSWS), and
10 particularly in the Pacific Northwest, although significant numbers may also occur in the
11 Upper Midwest (15-25% of population). In all cases, because the concentrations of both
12 nitrogen and phosphorus are low, additional inputs of nitrogen may have a limited potential to
13 cause eutrophication, because their input will quickly lead to a switch in the limiting nutrient;
14 additions of nitrogen to these systems would soon lead to nitrogen-sufficient and
15 phosphorus-deficient conditions. Increases in nitrogen deposition to some regions would
16 probably lead to measurable increases in algal biomass in lakes with both low concentrations
17 of dissolved nitrogen and substantial concentrations of phosphorus, but the number of lakes
18 that meet these criteria naturally (i.e., that do not have large anthropogenic inputs of
19 phosphorus) is likely to be quite small.
20 Few topics in aquatic biology have received as much attention in the past decade as the
21 debate over whether estuarine and coastal ecosystems are limited by nitrogen, phosphorus, or
22 some other factor (reviewed by Hecky and Kilham, 1988). Numerous geochemical and
23 experimental studies have suggested that nitrogen limitation is much more common in
24 estuarine and coastal waters than in freshwater systems. Experiments to confirm widespread
25 nitrogen limitation in estuaries have not been conducted, however, and nitrogen limitation
26 cannot be assumed to be the rule. Taken as a whole, the productivity of estuarine waters of
27 the United States correlates more closely with supply rates of nitrogen than of other nutrients
28 (Nixon and Pilson, 1983). Specific instances of phosphorus limitation (Smith, 1984) and of
29 seasonal switching between nitrogen and phosphorus limitation (D'Elia et al., 1986; McComb
30 et al., 1981) have been observed and stand as exceptions to the general rule of nitrogen
31 limitation in marine ecosystems. Nitrogen-fixing blue green algae are rarely abundant in
August 1991
10-236
DRAFT-DO NOT QUOTE OR CITE
-------
1 estuarine waters (Howarth et al., 1988a), and so nitrogen-deficient conditions may continue
2 indefinitely in these systems, unless nitrogen supply exceeds the biological demand for
3 nitrogen.
4 ' Estimation of the contribution of nitrogen deposition to the eutrophication of estuarine
5 and coastal waters is made difficult by the multiple direct anthropogenic sources (e.g., from
6 agriculture and sewage) of nitrogen against which the importance of atmospheric sources must
7 be weighed. Estuaries and coastal areas are natural locations for cities and ports, and most of
8 the watersheds of major estuaries in the United States have been substantially developed. The
9 crux of any assessment of the importance of nitrogen deposition to estuarine eutrophication is
10"'' establishing the relative importance of direct anthropogenic effects (e.g., sewage and
11 ' agricultural runoff) and indirect effects (e.g., atmospheric deposition). In the United States,
12 a large effort has been made to establish the relative importance of sources of nitrogen to the
13 Chesapeake Bay (e.g., D'Elia et al., 1982; Smullen et al., 1982; Fisher et al., 1988; Tyler,
14 1988). Estimates of the contribution of nitrogen to the Chesapeake Bay from each individual
15 source are very uncertain; estimating the proportion of nitrogen deposition exported from
16 forested watersheds is especially problematic but critical to the analysis, because ca. 80% of
17 the Chesapeake Bay basin is forested. Nonetheless, three attempts at determining the
18 proportion of the total NO3" load to the Bay attributable to nitrogen deposition all produce
19 estimates in the range of 18 to 31% (Table 10-35). Supplies of nitrogen from deposition
20 exceed supplies from all other non-point sources to the Bay (e.g., agricultural runoff,
21 pastureland runoff, urban runoff), and only point source inputs represent a greater input than
22 deposition.
23
24 10.7.8.3 Direct Toxicity
25 Toxic effects of nitrogen on aquatic biota result from un-ionized ammonia (NH3), which
26 occurs in equilibrium with ionized ammonium (NH4+) and OH". Ammonia concentrations
27 approach toxic concentrations most commonly at high pH and temperature values, which are
28 most typical of heavily polluted lakes and streams (e.g., Effler et al., 1990). In the well-
29 oxygenated conditions typical of unpolluted lakes and streams (as well as in most watersheds)
30 NH4+ is rapidly oxidized to NO3~, which does not have toxic effects on aquatic organisms.
31 Within the typical range of pH and temperature that unpolluted lakes and streams experience,
August 1991
10-237 DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-35. THREE NITROGEN BUDGETS FOR THE CHESAPEAKE BAY
Source of
Nitrogen
Direct Deposition
NO/
NH4
N Load to Bay (from direct
deposition)11
Forests
NO3" Deposition
NH4+ Deposition
Watershed Rentention
In-stream Retention
Atmospheric NO3' Load to
Bay (from forests)
N Load to Bay (from
forests)b
Pastureland
NO3" Deposition
NHL,* Deposition
Animal Wastes
Watershed Retention
In-stream Retention
Atmospheric NO3" Load to
Bay (from pastures)
N Load to Bay (from
pastures)1"
Cropland
NO3" Deposition
NIL,* Deposition
Fertilizers
Watershed Retention
In-stream Retention
Atmospheric NO3" Load to
Bay (from cropland)
N Load to Bay (from
cropland)11
Residential/Urban
NO3" Deposition
NIL,* Deposition
Watershed Retention
In-stream Retention
Atmospheric NO3' Load to
Bay (from urban areas)
N Load to Bay (from urban
areas)b
EDF
Budget
(eq X 109 • yf1
0.6
0.3
0.9
6.4
3.5
80%
50%
0.6
1.0
1.7
0.9
13.9
95 %c
50%°
0.5
1.1
1.8
., 1.0
11.3
70%
70%
0.6
i 4.2
0.3
0.2
35%
0%
0.2
0.3
Versar
Budget
(eq X 109 • yr-1
0.5
_a
0.5
6.0 ;
_a
95%
• 50%-..
0.15
0.15
1.2
'_« • • • '• •
8.4
94-99%
50%
0.01-0.04
0.05-0.3
2.0
a
5.9-19.3
76-99%
50%
0.0-0.2
0.04-2.6
0.5
-a
62-96%
20%
0.01-0.1
0.01-0.1
Refined
Budget
(eq X ,109 • yr-1
0.4
0.2
> 0.6
4-6
2.5
84.6%
0.5 35%
0.7
0.9
0.5
13.9
95 %d
35%
0.09
0.6
1.5
0.8
11.3
95%
35%
0.05
0.4
0.4
0.2
50%
35%
0.1
0.2
August 1991
10-238
DRAFT-DO NOT QUOTE OR CITE
-------
TABLE 10-35 (cont'd). THREE NITROGEN BUDGETS FOR THE
CHESAPEAKE BAY
Source of
Nitrogen
Point Sources
NO3- LOAD TO BAY
(FROM DEPOSITION
TOTAL NITROGEN LOAD
TO BAY"
% of N from NO3" depostion
EDF
Budget
(eq X 109 • yr-1
2.4
2.50
9.95
Versar
Budget
(eq X 109 • yfj
1.4-2.3
0.67-1.06
2.16-5.90
25%
Refined
Budget
(eq X 109 • yr'1
2.4 :
1.09
4.87
18-31%
22.5%
"The Versar Budget (Tyler, 1988) does not calculate loads of NH4+. ;
bFor the EDF Budget (Fisher et al., 1988a) and refined budget total nitrogen load to the Bay includes both NO3'
andNH4+. The Versar Budget (Tyler, 1988) includes only NO3".
"Watershed and In-stream retention values for pastureland in the EDF Budget apply only to animal wastes. For
atmospheric deposition, the cropland retention value (70%) was used. •
d95% retention was used for animal wastes; 85% retention was used for deposition (see text). .,
The range of contributions of NO3" deposition to the total budget were calculated by comparing maximum to
maximum estimates, and minimum to minimum estimates. These combinations are more likely to occur during
extreme (e.g., very wet or very dry) years.
1
2
3
4
5
6
7
toxic concentrations of NH3 .resulting from nitrogen deposition would be extremely unusual.
At a pH of 7, and a temperature of 15 °C, for example, concentrations of total NH4+ would
have to reach over 750 /imol • L"1 before chronically toxic concentrations of free NH3 would
develop. Currently no areas of North America are known to experience rates of nitrogen
deposition that are sufficient to produce such high concentrations of total NH4+ in surface
waters.
August 1991
10-239
DRAFT-DO NOT QUOTE OR CITE
-------
REFERENCES
Abd. Aziz, S. A.; Nedwell, D. B. (1986a) The nitrogen cycle of an east coast, U.K., saltmarsh: I. nitrogen
assimilation during primary production; detrital mineralization. Estuarine Coastal Shelf Sci. 22: 559-575.
Abd. Aziz, S. A.; NedweU, D. B. (1986b) The nitrogen cycle of an east coast, U.K. saltmarsh: II. nitrogen
fixation, nitrification, denitrification, tidal exchange. Estuarine Coastal Shelf Sci. 22: 689-704.
Abeles, F. B.; Craker, L. E.; Forrence, L. E.; Leather, G. R. (1971) Fate of air pollutants: removal of ethylene,
sulfur dioxide, and nitrogen dioxide by soil. Science (Washington, DC) 173: 914-916.
Aber, J. D.; Melillo, J. M. (1982) Nitrogen immobilization in decaying hardwood leaf litter as a function of
initial nitrogen and lignin content. Can. J. Bot. 60: .2263-2269.'
Aber, J. D.; Nadelhoffer, K. J.; Steudler, P.; Melillo, J. M. (1989) Nitrogen saturation in northern forest
ecosystems: excess nitrogen from fossil fuel combustion may stress the biosphere. Bioscience 39-
378-386.
Aerts, R.; Berendse, F. (1988) The effect of increased nutrient availability on vegetation dynamics in wet
heathlands. Vegetatio 76: 63-69.
Aerts, R.; Berendse, F. (1989) An analysis of competition in heathland ecosystems. L Competition for nutrients.
In: Aerts, R. Plant strategies and nutrient cycling in heathland ecosystems [Ph.D dissertation]. Utrecht,
The Netherlands: University of Utrecht; pp. 147-161.
Aerts, R.; de Caluwe, H.; Schmitz, M. (1989) An analysis of competition in heathland ecosystems. II.
Competition for light. la: Aerts, R. Plant strategies and nutrient cycling in heathland ecosystems [Ph.D
dissertation]. Utrecht, The Netherlands: University of Utrecht; pp. 165-180. !
Agren, G. I.; Bosatta, E. (1988) Nitrogen saturation of terrestrial ecosystems. Environ. Pollut. 54: 185-197.
i , ' •
Alexander, M. (1963) Soil microbiology. New York, NY: John Wiley and Sons.
Alexander, M. (1977) Introduction to soil microbiology. 2nd ed. New York, NY: John Wiley & Sons-
pp. 225-230; 239-246; 272-286; 355-356.
Altshuller, A. P.; Linthurst, R. A., eds. (1984) Acidic deposition phenomenon and its effects: critical assessment
review papers; volume I, atmospheric sciences; volume II, effects sciences. Washington, DC: U. S.
Environmental Protection Agency, Office of Research and Development; EPA report nos.
EPA/600/8-83/016aF-bF. Available from: NTIS, Springfield, VA; PB85-100030 and PB85-100048.
Altwicker, E. R.; Shanaghan, P. E.; Johannes, A. H. (1986) Atmospheric deposition to remote receptors:
HI. statistical analysis of precipitation data from the ILWAS-network. Water Air Soil Pollut. 28: 71-88.
Anderson, I. C.; Levine, J. S. (1987) Simultaneous field measurements of biogenic emissions of nitric oxide and
nitrous oxide. J. Geophys. Res. [Atmos.] 92: 965-976. , .
Aneja, V. P.; Rogers, H. H.; Stahel, E. P. (1986) Dry deposition of ammonia at environmental concentrations
on selected plant species. J. Air Pollut. Control Assoc. 36: 1338-1341.
Appel, B. R.; Tokiwa, Y. (1981) Atmospheric particulate nitrate sampling errors due to reactions with particulate
and gaseous strong acids. Atmos. Environ. 15: 1087-1089. •
August 1991
10-240 , DRAFT-DO NOT QUOTE OR CITE
-------
Armentano, T. V.; Menges, E. S. (1986) Patterns of change in the carbon balance of organic soil-wetlands of the
temperate zone. J. Ecol. 74: ,755-774. • •. :
Aronsson, A.; Elowson, S. (1980) Effects of irrigation and fertilization on mineral nutrients in Scots pine
needles. In: Persson, T., ed. Structure and function of northern coniferous forests - an ecosystem study.
Ecol. Bull. 32: 219-228. .
Aston, S. R. (1980) Nutrients, dissolved gases, and general biochemistry in estuaries. In: Olausson, E.; Cato, I.,
eds. Chemistry and biogeochemistry of estuaries. Chichester, United Kingdom: John Wiley and Sons;
pp. 233-262. • '
Atlas, R. M.; Bartha, R. (1981) Microbial ecology: fundamentals and applications. Reading, MA:
Addison-Wesley Publishing Company.
Bache, D. H. (1979a) Particle transport within plant canopies - I. a framework for analysis. Atmos. Environ.
13: 1257-1262.
Bache, D. H. (1979b) Particulate transport within plant canopies - II. prediction of deposition velocities. Atmos.
Environ. 13: 1681-1687.
Baird, S. F.; Buso, D. C.; Hornbeck, J. W. (1987) Acid pulses from snowmelt at acidic Cone Pond, New
Hampshire. Water Air Soil Pollut. 34: 325-338.
Baker, L. A. (1991) Regional estimates of atmospheric dry deposition. In: Charles, D. F., ed. Acidic deposition
and aquatic ecosystems: regional case 'studies. New York, NY: Springer-Verlag; Appendix C: hi press.
Baker, J. P.; Christensen, S. W. (1991) Effects of acidification on biological communities in aquatic ecosystems.
In: Charles, D. F., ed. Acidic deposition and aquatic ecosystems: regional case studies. New York, NY:
Springer-Verlag. •'.'•/.
Baker, L. A.; Kaufmann, P. R.; Herlihy, A. T.; Eilers, J. M. (n.d.) Current status of surface water acid-base
chemistry. National Acid Precipitation Assessment Program; state of the science/technology report 9: in
press. • •
Baker, J. B.; Switzer, G. L.; Nelson, L. E. (1974) Biomass production and nitrogen recovery after fertilization
of young loblolly pines. Soil Sci. Soc. Am. Proc. 38: 958-961.
Baker, T. G.; Oliver, G. R.; Hodgkiss, P. D. (1986) Distribution and cycling of nutrients in Pinus radiata as
affected by past lupin growth and fertiliser. For. Ecol. Manage. 17: 169-187.
Baker, J. P.; Freda, J.; Christensen, S. W.; Sale, M-, J- (1989) Biological effects of changes in surface water
.acid-base chemistry. Washington, DC: National Acid Precipitation Assessment Program;
state-of-science/technology report 13.
Baldocchi, D. (1988) A multi-layer model for estimating sulfur dioxide deposition to a deciduous oak forest
canopy. Atmos. Environ. 22: 869-884.
Baldocchi, D. D.; Hicks, B. ,B.; Camara, P. (1987) A canopy stomatal resistance model for gaseous deposition to
vegetated surfaces. Atmos. Environ. 21: 91-101.
Baldocchi, D. D.; Hicks, B. B.; Meyers, T..P. (1988) Measuring biosphere-atmosphere exchanges of biologically
related gases with micrometeorological methods. Ecology 69: 1331-1340.
August 1991
10-241
DRAFT-DO NOT QUOTE OR CITE
-------
Barbour, M. G.; Burk, J. H.; Pitts, W. D. (1980) Terrestrial plant ecology. 'Menlo Park, CA: The
Benjamin/Cummings Publishing Company, lac.; pp. 10-11,267-270: '
Barker, J. L.;'Witt, E. C. (1990) Effects of acidic precipitation on the water quality of streams in the Laurel Hill
area, Somerset County, Pennsylvania, 1983-1986. Harrisburg, PA: U. S. Geological Survey, Water
Resources Investigations; report 89-4113.
Barrie, L. A.; Sirois, A. (1986) Wet and dry deposition of sulphates and nitrates in eastern Canada: 1979-1982.
Water Air Soil. Pollut. 30: 303-310.
Bayley, S. E.; Vitt, D. H.; Newbury, R. W.; Beaty, K. G.; Behr, R.; Miller, C. (1987) Experimental
acidification of a Sphagnum-dominated peatland: first year results. Can. J. Fish. Aquat. Sci. 44(suppl. 1):
194-205. ._..,-...-
Begon, M.; Harper, J. L.; Townsend, C. R. (1986) Ecology: individuals; populations, and communities.
Sunderland, MA: Sinauef Associates, Inc.
Bennett, M. (1988) A simple physical model of dry deposition to a rough surface. Atmos. Environ.
22: 2701-2705.
Bennett, J. H.; Hill, A. C. (1973) Absorption of gaseous air pollutants by a standardized plant canopy. J. Air
Pollut. Control Assoc. 23: 203-206.
Berendse, F. (1985) The effect of grazing on the outcome of competition between plant species with different
nutrient requirements. Oikos 44: 35-39.
Berendse, F.; Aerts, R. (1984) Competition between Erica tetralix L. and Molinia caerulea (L.) Moench as
affected by the availability of nutrients. Acta Oebol. O'ecol. Plant. 5: 3-14.
Berger, M. G.; Klaus, R. E.; Fock, H. P. (1986) Assimilation of gaseous ammonia by sunflower leaves during
photosynthesis. Aust. J. Plant Physiol. 13: 211-219.
Betlach, M. R.; Tiedje, J. M. (1981) Kinetic explanation for accumulation of nitrite, nitric oxide, and nitrous
oxide during bacterial denitrification. Appl. Environ. Microbiol. 42: 1074-1084.
Biggar, C. M.; Cole, D. W. (1983) Effects of harvesting intensity on nutrient losses and future productivity in
high and low productivity red alder and Douglas-fir stands. In: Ballard, R.; Gessel, S. P., eds.
I.U.F.R.O. symposium on forest site and continuous productivity. Portland, OR: U. S. Forest Station,
Pacific Northwest Forest and Range Experiment Station; U. S. Forest Service general technical report no.
PNW-163.
Biggs, R. B.; Cronin, L. E. (1981) Special characteristics of estuaries. In: Neilson, B. J.; Cronin, L. E., eds.
Estuaries and nutrients. Clifton, NJ: Humana Press; pp. 3-23.
Bilby, R. E.; Likens, G. E. (1979) Effect of hydrologic fluctuations on the transport of fine particulate organic
carbon in a small stream. Limnol. Oceanogr. 24: 69-75.
Billings, W. D. (1978) Plants and the ecosystem. 3rd ed. Belmont, CA: Wadsworth Publishing Company, Inc.;
pp. 1-62, 83-108.
Binkley, D. (1986) Forest nutrition management. New York, NY: John Wiley & Sons.
Blackmer, A. M.; Bremner, J. M. (1978) Inhibitory effect of nitrate on reduction of N2O to N2 by soil
microorganisms. Soil Biol. Biochem. 10: 187-191.
August 1991
10-242 DRAFT-DO NOT QUOTE OR CITE
-------
Bockheim, J. G.; Leide, J.tE,; Tavella, D. S. (1986) Distribution and cycling of macronutrients in aPinus
resinosa plantation fertilized with nitrogen and.potassium. Can. J. For. Res, 16: 778-785.
Bolin, B.; Arrhenius, E., eds. (1977) Nitrogen - an essential life factor and a growing environmental hazard:
report from Nobel symposium no. 38. Ambio 6: 96-105.
Bolle, H.-J.; Seiler, W.; Bolin, B. (1986) Other greenhouse gases and aerosols: assessing their role for
atmospheric radiative transfer. In: Bolin, B.; Doos, B. R.; Jager, J.; Warrick, R. A., eds. The
greenhouse effect, climatic change, and ecosystems. SCOPE 29. Chichester, United Kingdom: John
Wiley and Sons; pp. 157-203.
Bondietti, E. A.; Baes, C. F., Ill; McLaughlin, S. B. (1989) Radial trends in cation ratios in tree rings as
indicators of the impact of atmospheric deposition on forests. Can. J. For. Res. 19: 586-594.
Boring, L. R.; Swank, W. T.; Waide, J. B.; Henderson, G, S. (1988) Sources, fates, and impacts of nitrogen
inputs to terrestrial ecosystems: review and synthesis. Biogeochemistry 6: 119-159.
Bormann, F. H. (1982) The effects of air pollution on the New England landscape. Ambio 11: 338-346.
Bormann, F. H.; Likens, G. E.; Melillo, J. M. (1977) Nitrogen budget for an aggrading northern hardwood
forest ecosystem. Science (Washington, DC) 196: 981-983.
Bosatta, E.; Berendse, F. (1984) Energy or nutrient regulation of decomposition: implications for the
mineralization-immobilization response to perturbations. Soil Biol. Biochem. 16: 63-6?.
Bosatta, E.; Staaf, H. (1982) The control of nitrogen turn-over in forest litter. Oikos 39: 143-151.
Boston, H. L. (1986) A discussion of the adaptations for carbon acquisition in relation to the growth strategy of
' aquatic isoetids. Aquat. Bot. 26: 259-270.
Bowden, W. B. (1986) Gaseous nitrogen emissions from undisturbed terrestrial ecosystems: an assessment of
their impacts on local and global nitrogen budgets. Biogeochemistry 2: 249-279.
Bowden, W. B. (1987) The biogeochemistry of nitrogen in freshwater wetlands. Biogeochemistry 4: 313-348.
Bowden, W. B.; Bormann, F. H. (1986) Transport and loss of nitrous oxide in soil water after forest
clear-cutting. Science (Washington, DC) 233: 867-869.
Bowersox, V. C.; Sisterson, D. L.; Olsen, A. R. (1990) Acid rain, a world-wide phenomenon: a perspective
from the United States. Int. J. Environ. Stud. 36: 83-101.
Boxman, D.; van Dijk, H.; Houdijk, A.; Roelofs, J. (1988> Critical loads for nitrogen with special emphasis on
ammonium. In: Nilsson, J.; Grennfelt, P., eds. Critical loads for sulphur and nitrogen: report from a
workshop; March; Skokloster, Sweden. Copenhagen, Denmark: Nordic Council of Ministers;
pp. 295-322.
Boyle, J. R.; Ek, A. R. (1972) An evaluation of some effects of bole and branch pulpwood harvesting on site
macronutrients. Can. J. For. Res. 2: 407-412.
Boyle, J. R.; Voigt, G. K. (1973) Biological weathering of silicate minerals: implications for tree nutrition and
soil genesis. Plant Soil 38:. 191-201.
August 1991
10-243 DRAFT-DO NOT QUOTE OR CITE
-------
Boynton, W. R.; Kemp, W. M.; Keefe, C. W. (1982) A comparative analysis of nutrients and other factors
influencing estuarine phytoplankton production. In: Estuarine comparisons. New York, NY: Academic
Press, Inc.; pp. 69-90.
Brakke, D. F.; Henriksen, A.; Norton, S. A. (1989) Estimated background concentrations of sulfate in dilute
lakes. Water Resour. Bull. 25: 247-253.
Brinson, M. M. (1977) Decomposition and nutrient exchange of litter in an alluvial swamp forest Ecology
58: 601-609.
Brinson, M. M.; Bradshaw, H. D.; Kane, E. S. (1984) Nutrient assimilative capacity of an alluvial floodplain
swamp. J. Appl. Ecol. 21: 1041-1057.
Broecker, W. S.; Peng, T.-H. (1982) Tracers in the sea. Eldigio.
Brown, D. J. A. (1988) Effect of atmospheric N deposition on surface water chemistry and the implications for
fisheries. Environ. Pollut. 54: 275-284.
Brown, K. A.; Freer-Smith, P. H.; Howells, G. D.; Skeffington, R. A.; Wilson, R. B. (1988) Rapporteurs'
report on discussions at the workshop on excess nitrogen deposition, Leatherhead, September 1987. In:
Dempster, J. P.; Manning, W. J.; Skeffington, R. A., eds. Excess nitrogen deposition: [papers from the
workshop]; September 1987; Leatherhead, Surrey, United Kingdom. Environ. Pollut. 54: 285-295.
Brunsting, A. M. H.; Heil, G. W. (1985) The role of nutrients in the interactions between a herbivorous beetle
and some competing plant species in heathlands. Oikos 44: 23-26.
Buckman, H. O.; Brady, N. C. (1969) The nature and property of soils. 7th. ed. New York, NY: The
Macmillan Company; pp. 437-462.
Buell, G. R.; Peters, N. E. (1988) Atmospheric deposition effects on the chemistry of a stream in northeastern
Georgia. Water Air Soil Pollut. 39: 275-291.
Buijsman, E. (1987) Ammonia emission calculation: fiction and reality. In: Asman, W. A. H.; Diederen,
H. S. M. A., eds. Ammonia and acidification: proceedings [of a] symposium of the European
Association for the Science of Air Pollution (EURASAP); April; Bilthoven, The Netherlands. European
Association for the Science of Air Pollution; pp. 13-27.
Buijsman, E.; Erisman, J.-W. (1987) Wet deposition of ammonium in Europe. In: Asman, W. A. H.; Diederen,
H. S. M. A., eds. Ammonia and acidification: proceedings of a symposium of the European Association
for the Science of Air Pollution (EURASAP); April; Bilthoven, The Netherlands. European Association
for the Science of Air Pollution; pp. 202-210.
Buresh, R. J.; Casselman, M. E.; Patrick, W. H., Jr. (1980) Nitrogen fixation in flooded soil systems, a review
Adv. Agron. 33: 149-192.
Buresh, R. J.; DeLaune, R. D.; Patrick, W. H., Jr. (1981) Influence of Spartina aherniflora on nitrogen loss
from marsh soil. Soil Sci. Soc. Am. J. 45: 660-661.
Businger, J. A. (1986) Evaluation of the accuracy with which dry deposition can be measured with current
micrometeorological techniques. J. dim. Appl. Meteorol. 25: 1100-1124.
Button, D. K. (1985) Kinetics of nutrient-limited transport and microbial growth. Microbiol. Rev. 49: 270-297.
August 1991
10-244 DRAFT-DO NOT QUOTE OR CITE
-------
Bytnerowicz, A.; Miller, P. R.; Olszyk, D. M. (1987) Dry deposition of nitrate, ammonium and sulfate to a
Ceanothus crassifolius canopy and surrogate surfaces. Atmos. Environ. 21: 1749-1757.
Bytnerowicz, A.; Miller, P. R.; Olszyk, D. M.; Dawson, P. L; Fox, C. A. (1987) Gaseous and particulate air
pollution in the San Gabriel Mountains of southern California. Atmos. Environ. 21: 1805-1814.
Cadle, S. H.; Countess, R. J.; Kelly, N. A. (1982) Nitric acid and ammonia in urban and rural locations.
Atmos. Environ. 16: 2501-2506.
Cadle, S. H.; Dasch, J. M.; Grossnickle, N. E. (1984) Retention and release of chemical species by a northern
Michigan snowpack. Water Air Soil Pollut. 22: 303-319.
Cadle, S. H.; Dasch, J. M.; Vande Kopple, R. (1987) Composition of snowmelt and runoff in northern
Michigan. Environ. Sci. Technol. 21: 295-299.
Cargill, S. M.; Jefferies, R. L. (1984) Nutrient limitation of primary production in a sub-arctic salt marsh.
J. Appl. Ecol. 21: 657-668.
Carpenter, E. J.; Capone, D. G., eds. (1983) Nitrogen in the marine environment. New York, NY: Academic
Press.
Caton, B. J. (1989) Nutrient reduction to the Bay: what's equitable, feasible, likely. In: Water-policy issues
related to the Chesapeake Bay. Virginia Water Resources Research Center; pp. 14-15.
Chatarpaul, L.; Robinson, J. B. (1979) Nitrogen transformations in stream sediments: 15N studies. In: Litchfield,
C. D.; Seyfried, P. L., eds. Methodology for biomass determinations and microbial activities in
sediments: a symposium sponsored by ASTM Committee D19 on Water; Ft. Lauderdale, FL; Jan. 1978.
Philadelphia, PA: American Society for Testing and Materials; pp. 119-125. (ASTM special technical
publication 673).
Chiandani, G.; Vighi, M. (1974) The N:P ratio and tests with Selanastrum to predict eiitrophication in lakes.
Water Res. 8: 1063-1069.
Chinchilli, J. E. (1989) Development and a clean bay: is coexistence possible? In: Water-policy issues related to
the Chesapeake Bay. Virginia Water Resources Research Center; pp. 6-9.
Cho, C. M.; Sakdinan, L. (1978) Mass spectrometric investigation on denitrification. Can. J. Soil Sci.
58: 443-457.
Clark, M. D.; Gilmour, J. T. (1983) The effect of temperature on decomposition at optimum and saturated soil
water'contents. Soil Sci. Soc. Am. J. 47: 927-929.
Code of Federal Regulations. (1987) Endangered and threatened wildlife and plants. C. F. R. 50: 17.11 & 17.12.
Cole, D. W.; Gessel, S. P. (1965) Movement of elements through a forest soil as influenced by tree removal and
fertilizer additions. In: Youngberg, C. T., ed. Forest-soil relationships in North America: papers
presented at the second North American forest soils conference; August 1963; Corvallis, OR: Oregon
State University Press; pp. 95-104.
Cole, D. W.; Rapp, M. (1981) Elemental cycling in forest ecosystems. In: Reichle, D. E., ed. Dynamic
properties of forest ecosystems. Cambridge, United Kingdom: Cambridge University Press; pp. 341-409.
(International biological programme 23).
August 1991
10-245 DRAFT-DO NOT QUOTE OR CITE
-------
Cole, D. W.; Crane, W. J. B.; Grier, C. C. (1975) The effect of forest management practices on water chemistry
in a second-growth Douglas fir ecosystem. In: Bernier, B.; Winget, C. H., eds. Forest soils and forest
land management. Quebec, Canada: Laval University Press; pp. 195-207.
Cole, J. J.; Howarth, R. W.; Nolan, S. S.; Marino, R. (1986) Sulfate inhibition of molybdate assimilation by
planktonic algae and bacteria: some implications for the aquatic nitrogen cycle. Biogeochemistry
2: 179-196.
Cook, E. R.; Johnson, A. H. (1989) Climate change and forest decline: a review of the red spruce case. Water
Air Soil Pollut. 48: 127-140.
Cooper, A. B. (1986) Supression of nitrate formation within an exotic conifer plantation. Plant Soil 93: 383-394.
Correll, D. L. (1981) Nutrient mass balances for the watershed, headwaters intertidal zone, and basin of the •
Rhode River estuary. Limnol. Oceanogr. 26: 1142-1149.
Correll, D. L.; Ford, D. (1982) Comparison of precipitation and land runoff as sources of estuarine nitrogen. '
Estuarine Coastal Shelf Sci. 15: 45-56.
Cosby, B. J.; Hornberger, G. M.; Galloway, J. N.; Wright, R. F. (1985) Modeling the effects of acid •
deposition: assessment of a lumped parameter model of soil water and streamwater chemistry. Water
Resour. Res. 21: 51-63.
Cosby, B. J.; Ryan, P. F.; Webb, R.; Hornberger, G. M.; Galloway, J. N. (1991) Mountains of West Virginia.
In: Charles, D. F., ed. Acidic deposition and aquatic ecosystems: regional case studies. New York, NY:
Springer-Verlag.
Cotrufo, C.; Berry, C. R. (1970) Some effects of a soluble NPK fertilizer on sensitivity of eastern white pine to
injury from SO2 air pollution. For. Sci. 16: 72-73.
Cowling, D. W.; Lockyer, D. R. (1981) Increased growth of ryegrass exposed to ammonia. Nature (London)
292: 337-338. •
Crutzen, P. J. (1970) The influence of nitrogen oxides on the atmospheric ozone content. Q. J. R Meteorol
Soc. 96: 320-325.
D'EHa, C. F. (1987) Too much of a good thing: nutrient enrichment of the Chesapeake Bay. Environment
29(2): 6-11, 30-33.
D'Elia, C. F.; Taft, J.; Smullen, J. T.; Macknis, J. (1982) Nutrient enrichment. In: Chesapeake Bay Program
technical studies: a synthesis. Annapolis, MD: U. S. Environmental Protection Agency; pp. 36-102.
Available from: NTIS, Springfield, VA; PB84-111202.
D'Elia, C. F.; Sanders, J. G.; Boynton, W. R. (1986) Nutrient enrichment studies in a coastal plain estuary:
phytoplankton growth in large-scale, continuous cultures. Can. J. Fish. Aquat. Sci. 43: 397-406.
Dabney, S. M.; Bouldin, D. R. (1985) Fluxes of ammonia over an alfalfa field. Agron. J. 77: 572-578.
Darnman, A. W. H. (1988) Regulation of nitrogen removal and retention in Sphagnum bogs and other peatlands.
Oikos 51: 291-305.
Dasch, J. M. (1987) Measurement of dry deposition to surfaces in deciduous and pine canopies. Environ. Pollut
44:261-277.
August 1991
10-246 DRAFT-DO NOT QUOTE OR CITE
-------
Dasch, J. M. (1989) Dry deposition of sulfur dioxide or nitric acid to oak, elm and pine leaves. Environ. Pollut.
' 59: 1-16.
Davidson, E. A.; Swank, W. T. (1987) Factors limiting denitrification in soils from mature and disturbed
southeastern hardwood forests. For. Sci. 33: 135-144.
Davidson, E. A.; Swank, W. T. (1990) Nitrous oxide dissolved in soil solution: an insignificant pathway of
nitrogen loss from a southeastern hardwood forest. Water Resour. Res. 26: 1687-1690.
Davidson, C. I.; Wu, Y. -L. (1990) Dry deposition of particles and vapors. In: Lindberg, S. E.; Page, A. L.;
Norton, S. A., eds. Acidic precipitation: v. 3, sources, deposition, and canopy interactions: -New York,
,...; NY: Springer-Verlag; pp. 103-216. - .
Davidson, E. A.; Stark, J. M.; Firestone, M. K. (1990) Microbial production and consumption of nitrate in an
annual grassland. Ecology 71: 1968-1975.
Davis, C. B.; van der Valk, A. G.; Baker, J. L. (1983) The role of four macrophyte species in the removal of
nitrogen and phosphorus from nutrient-rich water in a prairie marsh, Iowa. Madrono 30: 133-142.
Day, J. W., Jr.; Hopkinson, C. S.; Conner, W. H. (1982) An analysis of environmental factors regulating
community metabolism and fisheries production in a Louisiana estuary. In: Kennedy, V. S., ed. Estuarine
comparisons. New York, NY: Academic Press; pp. 121-136.
De,Temmerman, L.; Ronse, A.; van den Cruys, K.; Meeus-Verdinne, K. (1988) Ammonia and pine tree dieback
in Belgium. In: Mathy, P., ed. Air pollution and ecosystems: proceedings of an international symposium;
May 1987; Grenoble, France. Boston, MA: D. Reidel Publishing Company; pp. 774-779.
de Vries,- W. (1988) Critical deposition levels for nitrogen and sulphur on Dutch forest ecosystems. Water Air
Soil Pollut. 42: 221-239.
Dean, J. V.; Biesboer, D. D. (1985) Loss and uptake of 15N-ammonium in submerged soils of a cattail marsh.
Am. J. Bot. 72: 1197-1203.
Dean, J. V.; Harper, J. E. (1986) Nitric oxide and nitrous oxide production by soybean and winged bean during
the in vivo nitrate reductase assay. Plant Physiol. 82: 718-723.
Dean, J. V.; Harper, J. E. (1988) The conversion of nitrite to nitrogen oxide(s) by the constitutive
NAD(P)H-nitrate reductase enzyme from soybean. Plant Physiol. 88: 389-395.
DeBusk, W. F.; Reddy, K. R. (1987) Removal of flopdwater nitrogen in a cypress swamp receiving primary
wastewater effluent. Hydrobiologia 153: 79-86.
DeHayes, D. H.; Ingle, M. A.; Waite, C. E. (1989) Nitrogen, fertilization enhances cold tolerance of red spruce
seedlings. Can.,J. For. Res. 19: 1037-1043.
Delany, A. C.; Davies, T. D. (1983) Dry deposition of NOX to grass in rural East Anglia. Atmos. Environ.
17: 1391-1394.
Delany, A. C.; Fitzjarrald, D. R.; Lenschow, D. H.; Pearson, R., Jr.; Wendel, G. J.; Woodruff, B. (1986)
Direct measurements of nitrogen oxides and ozone fluxes over grassland. J. Atmos. Chem. 4: 429-444.
DeLaune, R. D.; Reddy, C. N.; Patrick, W. H., Jr. (1981) .Organic matter decomposition in soil as influenced
by pH and redox conditions. Soil Biol. Biochem. 13: 533-534.
. August 1991
10_247 DRAFT-DO NOT QUOTE OR CITE
-------
DeLaune, R. D.; Smith, C. J.; Patrick, W. H., Jr. (1983) Nitrogen losses from a Louisiana Gulf Coast salt
marsh. Estuarine Coastal Shelf Sci. 17: 133-141.
DeLaune, R. D.; Smith, C. J.; Sarafyan, M. N. (1986) Nitrogen cycling in a freshwater marsh of Panicum
hemitomon on the deltaic plain of the Mississippi River. J. Ecol. 74: 249-256.
Delwiche, C. C. (1970) The nitrogen cycle. Sci. Am. 223: 137-147.
Delwiche, C. C. (1977) Energy relations in the global nitrogen cycle. Ambio 6: 106-111.
Denmead, O. T.; Simpson, J. R.; Freney, J. R. (1974) Ammonia flux into the atmosphere from a grazed pasture.
Science (Washington, DC) 185: 609-610.
Denmead, O. T.; Freney, J. R.; Simpson, J. R. (1976) A closed ammonia cycle within a plant canopy. Soil Biol
Biochem. 8: 161-164.
Denmead, O. T.; Nulsen, R.; Thurtell, G. W. (1978) Ammonia exchange over a corn crop. Soil Sci. Soc
Am. J. 42: 840-842.
DeWalle, D. R.; Sharpe, W. E.; Edwards, P. J. (1988) Biogeochemistry of two Appalachian deciduous forest
sites in relation to episodic stream acidification. Water Air Soil Pollut. 40: 143-156.
Dicker, H. J.; Smith, D. W. (1980) Physiological ecology of acetylene reduction (nitrogen fixation) in a
Delaware salt marsh. Microb. Ecol. 6: 161-171.
Dierberg, F. E.; Brezonik, P. L. (1982) Nitrifying population densities and inhibition of ammonium oxidation in
natural and sewage-enriched cypress swamps. Water Res. 16: 123-126.
Dillon, P. J.; Rigler, F. H. (1974) The phosphorus-chlorophyll relationship in lakes. Limnol. Oceanogr.
19: 767-773.
Dillon, P. J.; Lusis, M.; Reid, R.; Yap, D. (1988) Ten-year trends in sulphate, nitrate and hydrogen deposition
in central Ontario. Atmos. Environ. 22: 901-905.
Dodds, W. K.; Priscu, J. C. (1990) A comparison of methods for assessment of nutrient deficiency of
phytoplankton in a large oligotrophic lake. Can. J. Fish. Aquat. Sci. 47: 2328-2338.
Doering, P. H.; Oviatt, C. A.; Beatty, L. L.; Banzon, V. F.; Rice, R.; Kelly, S. P.; Sullivan, B. K.; Frithsen,
J. B. (1989) Structure and function in a model coastal ecosystem: silicon, the benthos and eutrophication.
Mar. Ecol. Prog. Ser. 52: 287-299.
Dollard, G. J.; Atkins, D. H. F.; Davies, T. J.; Healy, C. (1987) Concentrations and dry deposition velocities
of nitric acid. Nature (London) 326: 481-483.
Dolske, D. A. (1988) Dry deposition of airborne sulfate and nitrate to soybeans. Environ. Pollut. 53: 1-12.
Draaijers, G. P. J.; Ivens, W. P. M. F.; Bos, M. M.; Bleuten, W. (1989) The contribution of ammonia
emissions from agriculture to the deposition of acidifying and eutrophying compounds onto forests.
Environ. Pollut. 60: 55-66.
Driscoll, C. T.; Newton, R. M. (1985) Chemical characteristics of Adirondack lakes. Environ. Sci. Technol
19: 1018-1024.
August 1991
10-248 DRAFT-DO NOT QUOTE OR CITE
-------
Driscoll C. T.; Schaefer, D. A. (1989) Background on nitrogen processes. In: Malanchuk, J, L.; Nilsson, J.,
eds. The role of nitrogen in the acidification of soils and surface waters. Miljorapport, Sweden: Nordic
Council of Ministers; pp. 4-1 - 4-12.
Driscoll, C. T.; Schafran, G. C. (1984) Short-term changes in the base neutralizing capacity of an acid
Adirondack lake, New York. Nature (London) 310: 308-310.
Driscoll, C. T. et al. (n.d.) Processes regulating seasonal variations in the water chemistry of Adirondack lakes.
Water Air Soil Pollut.: in review.
Driscoll C. T.; Wyskowski, B. J.; Cosentini, C. G.; Smith, M. E. (1987a) Processes regulating temporal
longitudinal variations in the chemistry of a low-order woodland stream in the Adirondack region of New
York. Biogeochemistry 3: 225-241.
Driscoll, C. T.; Yatsko, C. P.; Unangst, F. J. (1987b) Longitudinal and temporal trends in the water chemistry
of the north branch of the Moose River. Biogeochemistry 3: 37-61.
Driscoll C. T.; Schaefer, D. A.; Molot, L. A.; Dillon, P. J. (1989a) Summary of North American data. In:
Malanchuk, J. L.; Nilsson, J., eds. The role of nitrogen in the acidification of soils and surface waters.
Gotab, Sweden: Nordic Council of Ministers; pp. 6-1 - 6-45.
Driscoll C T ; Wyskowski, B. J.; DeStaffan, P. (1989b) Chemistry and transfer of aluminum in a forested
watershed in the Adirondack region of New York, USA. In: Lewis, T., ed. The environmental chemistry
of aluminum. Chelsea, MA: Lewis Publishers; pp. 1-26.
Driscoll, C. T.; Newton, R. M.; Gubala, C. P.; Baker, J. P.; Christensen, S. W. (1991) Adirondack mountains.
In: Charles, D. F., ed. Acidic deposition and aquatic ecosystems: regional case studies. New York, NY:
Springer-Verlag; pp. 133-202.
Duff, J. H.; Triska, F. J. (1990) Denitrification in sediments from the hyporheic zone adjacent to a small
forested stream. Can. J. Fish. Aquat. Sci. 47: 1140-1147.
Duyzer, J. H.; Meyer, G. M.; van Aalst, R. M. (1983) Measurement of dry deposition velocities of NO, NO2
'and O3 and the influence of chemical reactions. Atmos. Environ. 17: 2117-2120.
Duyzer, J. H.; Bouman, A. M. M.; van Aalst, R. M.; Diederen, H. S. M. A. (1987) Assessment of dry
deposition fluxes of NH3 and NH4+ over natural terrains. In: Asman, W. A. H.; Diederen, H. S. M. A.,
eds Ammonia and acidification: proceedings of a symposium of the European Association for the Science
of Air Pollution (EURASAP); April; Bilthoven, The Netherlands. European Association for the Science
of Air Pollution; pp. 97-106. '
Edwards, P. J.; Helvey, J. D. (1991) Long-term ionic increases from a central Appalachian forested watershed.
J. Environ. Qual. 20: 250-255.
Effler, S. W.; Brooks, C. M.; Auer, M. T.; Doerr, S. M. (1990) Free ammonia and toxicity criteria in a
polluted urban lake. Res. J. Water Pollut. Control Fed. 62: 771-779.
Eilers, J. M.; Kanciruk, P.; McCord, R. A.; Overton, W. S.; Hook, L.; Blick, D. J.; Brakke, D. F.; Kellar,
P E.; DeHaan, M. S.; Silverstein, M. E.; Landers, D. H. (1987) Western lake survey phase I:
characteristics of lakes in the western United States, volume II: data compendium for selected physical
and chemical variables. Washington, DC: U. S. Environmental Protection Agency, Office of Acid
Deposition, Environmental Monitoring and Quality Assurance; EPA report no. EPA/600/3-86/054B.
Available from: NTIS, Springfield, VA; PB88-146832/REB.
August 1991
10-249 DRAFT-DO NOT QUOTE OR CITE
-------
Eilers, J. M.; Sullivan, T. J.; Hurley, K. C. (1990) The most dilute lake in the world? Hydrobiologia 199: 1-6.
Elkiey, T.; Ormrod, D. P. (1981) Absorption of ozone, sulphur dioxide, and nitrogen dioxide by petunia plants.
Environ. Exp. Bot. 21: 63-70.
Elkiey, T.; Ormrod, D. P.; Marie, B. (1982) Foliar sorption of sulfur dioxide, nitrogen dioxide, and ozone by
ornamental woody plants. Hortscience 17: 358-360.
Elkins, J. W.; Wofsy, S. C.; McElroy, M. B.; Kolb, C. E.; Kaplan, W. A. (1978) Aquatic sources and sinks
for nitrous oxide. Nature (London) 275: 602-606.
Ellenberg, H. (1987) Floristic changes due to eutrophication. In: Asman, W. A. H.; Diederen, S. M. A., eds.
Ammonia and acidification: proceedings of a symposium of the European Association for the Science of
Air Pollution (EURASAP); April; Bilthoven, The Netherlands. European Association for the Science of
Air Pollution; pp. 301-308.
Ellenberg, H. (1988) Floristic changes due to nitrogen deposition in central Europe, m: Nilsson, J.;
Grennfelt, P., eds. Critical loads for sulphur and nitrogen: report from a workshop; March; Skokloster,
Sweden. Copenhagen, Denmark: Nordic Council of Ministers; pp. 375-383.
Elser, J. J.; Marzolf, E. R.; Goldman, C. R. (1990) Phosphorus and nitrogen limitation of phytoplankton growth
in the freshwaters of North America: a review and critique of experimental enrichments. Can. J. Fish
Aquat. Sci. 47: 1468-1477.
Elwood, J. W.; Turner, R. R. (1989) Streams: water chemistry and ecology. In: Johnson, D. W.; Van Hook,
R. I., eds. Analysis of biogeochemical cycling processes in Walker Branch Watershed. New York, NY:
Springer-Verlag; pp. 301-350.
Elwood, J. W.; Newbold, J. D.; O'Neill, R. V. (1980) Phosphorus spiralling in a woodland stream ecosystem:
seasonal changes and the role of the microbial community. Int. Assoc. Theor. Appl. Limnol. (Tokyo).
Elwood, J. W.; Sale, M. J.; Kaufmann, P. R.; Cada, G. F. (1991) The Southern Blue Ridge province. In:
Charles, D. F., ed. Acidic deposition and aquatic ecosystems: regional case studies. New York, NY:
Springer-Verlag.
Emerson, K.; Russo, R. C.; Lund, R. W.; Thurston, R. V. (1975) Aqueous ammonia equilibrium calculations:
effect of pH and temperature. J. Fish. Res. Board Can. 32: 2379-2383.
Enders, G.; Teichmann, U. (1986) GASDEP - gaseous deposition measurements of SO2, NOX, and O3 to a spruce
stand: conception, instrumentation, and first results of an experimental project. In: Georgii, H.-W., ed.
Atmospheric pollutants in forest areas: their deposition and interception. Boston, MA: D. Reidel
Publishing Company; pp. 13-24.
Enoksson, V.; Sorensson, F.; Graneli, W. (1990) Nitrogen transformations in the Kattegat. Ambio 19: 159-166.
Eppley, R. W.; Renger, E. H.; Harrison, W. G.; Cullen, J. J. (1979) Ammonium distribution in southern
California coastal waters and its role in the growth of phytoplankton. Lirnnol. Oceanogr, 24: 495-509.
Erisman, J.-W.; Vermetten, A. W. M.; Asman, W. A. H.; Waijers-Upelaan, A.; Slanina, J. (1988) Vertical
distribution of gases and aerosols: the behaviour of ammonia and related components in the lower
atmosphere. Atmos. Environ. 22: 1153-1160.
Eshleman, K. N. (1988) Predicting regional episodic acidification of surface waters using empirical models.
Water Resour. Res. 24: 1118-1126.
August 1991
10-250 DRAFT-DO NOT QUOTE OR CITE
-------
Eshleman, K. N.; Hemond, H. F. (1985) The role of organic acids in the acid-base status of surface waters at
Bickford Watershed, Massachusetts. Water Resour. Res. 21: 1503-1510.
Evans, L. S. (1986) Proposed mechanisms of initial injury-causing apical dieback in red spruce at high elevation
in eastern North America. Can. J. For. Res. 16: 1113-1116.
Fahey, D. W.; Hubler, G.; Parrish, D. D.; Williams, E. J.; Norton, R. B.; Ridley, B. A.; Singh, H. B.; Liu,
S. C.; Fehsenfeld, F. C. (1986) Reactive nitrogen species in the troposphere: measurements of NO, NO2,
HNO3, particulate nitrate, peroxyacetyl nitrate (PAN), O3, and total reactive odd nitrogen (NOy) at Niwot
Ridge, Colorado. J. Geophys. Res. [Atmos.] 91: 9781-9793.
Farquhar, G. D.; Wetselaar, R.; Firth, P. M. (1979) Ammonia volatilization from senescing leaves of maize.
Science (Washington, DC) 203: 1257-1258.
Farquhar, G. P.; Firth, P. M.; Wetselaar, R.; Weir, B. (1980) On the gaseous exchange of ammonia between
leaves and the environment: determination of the ammonia compensation point. Plant Physiol.
66:710-714. ,
Farquhar, G. D.; Wetselaar, R.; Weir, B. (1983) Gaseous nitrogen losses from plants. In: Freney, J. R.;
Simpson, J. R., eds. Gaseous loss of nitrogen from plant-soil systems. The Hague, The Netherlands:
Martinus Nijhoff/Dr W. Junk Publishers; pp. 159-180.
Federal Register. (1985) Endangered and threatened wildlife and plants; review of plant taxa for listing as
endangered or threatened species. F. R. (September 27) 50: 39526-39584.
Federer, C. A.; Hombeck, J. W.; Tritton, L. M.; Martin, C, W.; Pierce, R. S. (1989) Long-term depletion of
calcium and other nutrients in eastern US forests Environ. Manage. 13: 593-601.
Feller, M. C. (1987) The influence of acid precipitation on stream chemistry in a small forested basin in
southwestern British Columbia. In: Swanson, R. H.; Bernier, P. Y.; Woodward, P. D., eds. Forest
hydrology and watershed management: proceedings of the Vancouver symposium; August. IAHS-AISH
publication no. 167. ,
Ferguson, P.; Lee, J. A. (1980) Some effects of bisulphite and sulphate on the growth of Sphagnum species in
the field. Environ. Pollut. Ser. A 21: 59-71.
Ferguson, P.; Robinson, R. N.; Press, M. C.; Lee, J. A. (1984) Element concentrations in five Sphagnum
species in relation to atmospheric pollution. J. Bryol. 13: 107-114.
Fisher, D.; Ceraso, J.; Mathew, T.; Oppenheimer, M. (1988a) Polluted coastal waters: the role of acid rain.
New York, NY: Environmental Defense Fund. ,
Fisher, T. R.; Harding, L. W., Jr.; Stanley, D. W.; Ward, L. G. (1988b) Phytoplankton, nutrients, and
turbidity in the Chesapeake, Delaware, and Hudson estuaries. Estuarine Coastal Shelf Sci. 27: 61-93.
Fisher, F. M.; Zak, J. C.; Cunningham, G. L.; Whitford, W. G. (1988c) Water and nitrogen effects on growth
and allocation patterns of creosotebush in the northern Chihuahuan Desert. J. Range Manage.
41: 387-391.
Fitzjarrald, D. R.; Lenschow, D. H. (1983) Mean concentration and flux profiles for chemically reactive species
in the atmospheric surface layer. Atmos. Environ. 17: 2505-2512.
Fleischer, S.; Stibe, L. (1989) Agriculture kills marine fish in the 1980s. Who is responsible for fish kills in the
year 2000? Ambio 18: 347-350.
August 1991
10-251 DRAFT-DO NOT QUOTE OR CITE
-------
Flett, R. J.; Schindler, D. W.; Hamilton, R. D.; Campbell, N. E. R. (1980) Nitrogen fixation in Canadian
Precambrian Shield lakes. Can. J. Fish. Aquat. Sci. 37: 494-505.
Flora, M. D.; Rosendahl, P. C. (1982) The impact of atmospheric deposition on the water quality of Everglades
National Park. In: International symposium on hydrometeorology; June. American Water Resources
Association; pp. 55-61.
Focht, D. D. (1974) The effect of temperature, pH, and aeration on the production of nitrous oxide and gaseous
nitrogen - zero-order kinetic model. Soil Sci. 118: 173-179.
Foerstel, H.; Latus, C.; Neubert, A. (1989) Einbau von Stickstoff aus gasformigem 15NO2 in die oberirdische
Biomasse von Pflanzen. In: Volume of poster abstracts: international congress on forest decline research -
state of knowledge and perspectives. Friedrichschafen, Federal Republic of Germany; pp. 353-354.
Foster, N. W. (1985) Acid precipitation and soil solution chemistry within a maple-birch forest in Canada. For.
Ecol. Manage. 12: 215-231.
Foster, N. W.; Nicolson, J. A. (1988) Acid deposition and nutrient leaching from deciduous vegetation and
podzolic soils at the Turkey Lakes Watershed. Can. J. Fish. Aquat. Sci. 45(suppl. 1): 96-100.
Foster, N. W.; Beauchamp, E. G.; Corke, C. T. (1985a) Reactions of 15N-labelled urea with jack pine
forest-floor materials. Soil Biol. Biochem. 17: 699-703.
Foster, N. W.; Beauchamp, E. G.; Corke, C. T. (1985b) Immobilization of nitrogen-15-labelled urea in a jack
pine forest floor. Soil Sci. Soc. Am. J. 49: 448-452.
Foster, N. W.; Nicolson, J. A.; Hazlett, P. W. (1989) Temporal variation in nitrate and nutrient cations in
drainage waters from a deciduous forest. J. Environ. Qual. 18: 238-244.
Fowler, D. (1984) Transfer to terrestrial surfaces. Philos. Trans. R. Soc. London B 305: 281-297.
Fowler, D.; Cape, J. N.; Unsworth, M, H. (1989a) Deposition of atmospheric pollutants on forests. Philos.
Trans. R. Soc. London B 324: 247-265.
Fowler, D.; Cape, J. N.; Deans, J. D.; Leath, I. D.; Murray, M. B.; Smith, R. I.; Sheppard, L. J.; Unsworth,
M. H. (1989b) Effects of acid mist on the frost hardiness of red spruce seedlings. New Phytol
106: 225-227.
Fox, D. G.; Barthuska, A. M.; Byrne, J. G.; Cowling, E.; Fisher, R.; Likens, G. E.; Lindberg, S. E.;
Linthurst, R. A.; Messer, J.; Nichols, D. S. (1989) A screening procedure to evaluate air pollution
effects on Class I wilderness areas. Ft. Collins, CO: U. S. Department of Agriculture Forest Service,
Rocky Mountain Forest and Range Experiment Station; general technical report RM-168.
Franklin, J. F.; Dyrness, C. T.; Moore, D. G.; Tarrant, R. F. (1968) Chemical soil properties under coastal
Oregon stands of alder and conifers. In: Trappe, J. M.; Franklin, J. F.; Tarrant, R. F.; Hansen, G. M.,
eds. Biology of alder: proceedings of a symposium; April 1967; Pullman, WA. Portland, OR: U. S.
Forest Service, Pacific Northwest and Range Experiment Station; pp. 157-172.
Prayer, W. E.; Monahan, T. J.; Dowden, D. C.; Graybill, F. A. (1983) Status and trends of wetlands and
deepwater habitats in the conterminous United States, 1950s to 1970s. Ft. Collins, CO: Colorado State
University, Department of Forest and Wood Sciences.
Freer-Smith, P. H. (1983) Chronic pollution injury to some tree species in response to SO2 and NO2 mixtures
[Thesis]. Lancaster, United Kingdom: University of Lancaster.
August 1991
10-252 DRAFT-DO NOT QUOTE OR CITE
-------
Freer-Smith, P. H. (1988) Direct effects 'of dry and wet deposition on forest ecosystems - in particular canopy
interactions. In: Mathy, P., eds. Air pollution and ecosystems. Dordrecht, The Netherlands: D. Reidel
Publishing Company; pp. 217-274.
Friedland, A. J.; Gregory, R. A.; Karenlampi, L.; Johnson, A. H. (1984) Winter damage to foliage as a factor
in red spruce decline. Can. J. For. Res. 14: 963-965.
Fuhrer, J.; Erismann, K. H. (1980) Uptake of NO2 by plants grown at different salinity levels. Experientia
36: 409-410. .
Gallagher, J. L. (1975) Effect of an ammonium nitrate pulse on the growth and elemental composition of natural
stands of Spartina alterniflofa and Juncus roemerianus. Am. J. Bot. 62: 644-648.
Gallagher, J. L.; Donovan, L. A.; Grant, D. M.; Decker, D. M. (1987) Interspecific differences in dead plant
buffering capacity alter the impact of acid rain on decomposition rates in tidal marshes. Water Air Soil
Pollut. 34: 339-346.
Galloway, J. N.; Schofield, C. L.; Hendrey, G. R.; Peters, N. E.; Johannes, A. H. (1980) Sources of acidity in
three lakes acidified during snowmelt. In: Drablos, D.; Tollan, A., eds. Ecological effects of acid
precipitation: proceedings of an international conference. Mysen, Norway: Johs. Grefslie Trykkeri A/S;
pp. 264-265.
Galloway, J. N.; Likens, G. E.; Keene, W. C.; Miller, J. M. (1982) The composition of precipitation in remote
areas of the world. J. Geophys. Res. C: Oceans Atmos. 87: 8771-8786.
Galloway, J. N.; Norton, S. A.; Church, M. R. (1983) Freshwater acidification from atmospheric deposition of
sulfuric acid: a conceptual model. Environ. Sci. Technol. 17: 541A-545A.
Garner, J. H. B.; Pagano, T'.; Cowling, E. B. (1989) An evaluation of the role of ozone, acid deposition, and
other airborne pollutants hi the forests of eastern North America. Asheville, NC: U. S. Department of
Agriculture, Forest Service, Southeastern Forest Experiment Station; general technical report SE-59.
Garten, C. T., Jr.; Hanson, P. J. (1990) Foliar retention of 15N-nitrate and 15N-ammonium by red maple (Acer
rubruni) and white oak (Quercus alba) leaves from simulated rain. Environ. Exp. Bot. 30: 333-342.
Ga'vis, J. (1976) Munk and Riley revisited: nutrient diffusion transport and rates of phytoplankton growth.
J. Mar. Res. 34: 161-179. ,
Gerhart, D. Z.; Likens, G. E. (1975) Enrichment experiments for determining nutrient limitation: four methods
compared. Limnol. Oceanogr. 20: 649-653.
• Gessel, S. P.; Cole, D. W.; Steinbrenner, E. C. (1973) Nitrogen balances in forest ecosystems of the Pacific
Northwest. Soil Biol. Biochem. 5: 19-34.
Gherini, S. A.; Mok, L.; Hudson, R. J. M.; Davis, G. F.; Chen, C. W.; Goldstein, R. A. (1985) The ILWAS
model: formulation and application. Water Air Soil Pollut. 26: 425-459.
Ghiorse, W. C.; Alexander, M. (1976) Effect of microorganisms on the sorption of sulfur dioxide and nitrogen
dioxide in soil. J. Environ. Qual. 5: 227-230.
Gilbert, D. A.; Sagraves, T. H.; Lang, M. M.; Mimson, R. K.; Gherini, S. A. (1989) R&D Lake Acidification
Assessment Project: Blue Lake acidification study. San Ramon, CA: Pacific Gas and Electric; report
009.5-89.2.
August 1991
10-253
DRAFT-DO NOT QUOTE OR CITE
-------
Gmur, N. F.; Evans, L. S.; Cunningham, E. A. (1983) Effects of ammonium sulfate aerosols on vegetation—
II. mode of entry and responses of vegetation. Atmos. Environ. 17: 715-721.
Godshalk, G. L.; Wetzel, R. G. (1978) Decomposition of aquatic angiosperms. .II. Particulate components.
Aquat. Bot. 5: 301-327.
Goldman, J. C. (1976) Identification of nitrogen as a growth-limiting nutrient in wastewaters and coastal marine
waters through continuous culture algal assays. Water Res. 10: 97-104.
Goldman, C. R. (1988) Primary productivity, nutrients, and transparency during the early onset of eutrophication
in ultra-oligotrophic Lake Tahoe, California-Nevada. Limnol. Oceanogr. 33: 1321-1333.
Goldman, J. C.; Gilbert, P. M. (1982) Comparative rapid ammonium uptake by four species of marine
phytoplankton. Limnol. Oceanogr. 27: 814-827.
Goodroad, L. L.; Keeney, D. R. (1984) Nitrous oxide emission from forest, marsh, and prairie ecosystems.
J. Environ. Qual. 13: 448-452.
Gordon, A. S.; Cooper, W. J.; Scheidt, D. J. (1986) Denitrification in marl and peat sediments in the Florida
Everglades. Appl. Environ. Microbiol. 52: 987-991.
Granat, L.; Johansson, C. (1983) Dry deposition of SO2 and NOX in winter. Atmos. Environ. 17: 191-192.
Graneli, E. (1978) Algal assay of limiting nutrients for phytoplankton production in the Oresund. Vatten
2: 117-128.
Graneli, E.; Wallstrom, K.; Larsson, U.; Graneli, W.; Elmgren, R. (1990) Nutrient limitation of primary
production in the Baltic Sea area. Ambio 19: 142-151.
Gravenhorst, G.; Hoefken, K. D.; Georgii, H. W. (1983) Acidic input to a beech and spruce forest. In:
Beilke, S.; Elshout, A. J., eds, Acid deposition: proceedings of the CEC workshop organized as part of
the concerted action "Physico-chemical behaviour of atmospheric pollutants"; September 1982; Berlin,
Federal Republic of Germany. Dordrecht, The Netherlands: D. Reidel Publishing Company;
pp. 155-171.
Grennfelt, P.; Hultberg, H. (1986) Effects of nitrogen deposition on the acidification of terrestrial and aquatic
ecosystems. Water Air Soil Pollut. 30: 945-963.
Grennfelt, P.; Bengtson, C.; Skarby, L. (1983) Dry deposition of nitrogen dioxide to Scots pine needles. In:
Pruppacher, H. R.; Semonin, R. G.; Slinn, W. G. N.j eds. Precipitation scavenging, dry deposition, and
resuspension. Volume 2. Dry deposition and resuspension: proceedings of the fourth international
conference; November-December 1982; Santa Monica, CA. New York, NY: Elsevier; pp. 753-762.
Grier, C. C. (1975) Wildfire effects on nutrient distribution and leaching in a coniferous ecosystem. Can. J. For.
Res. 5: 599-607.
Grime, J. P. (1979) Plant strategies and vegetation processes. Chichester: John Wiley & Sons.
Groffman, P. M.; House, G. J.; Hendrix, P. F.; Scott, D. E.; Crossley, D. A., Jr. (1986) Nitrogen cycling as
affected by interactions of components in a Georgia piedmont agroecosystem. Ecology 67: 80-87.
August 1991
10-254 DRAFT-DO NOT QUOTE OR CITE
-------
Guderian, R.; Kueppera, K. (1980) Response of plant communities to air pollution. In: Miller, P. R., ed.
Proceedings of the symposium on effects of 'air pollutants on Mediterranean and temperate forest
ecosystems; June; Riverside, CA. Berkeley, CA: U. S. Department of Agriculture, Pacific Southwest
Forest and Range Experiment Station; pp. 187-199; Forest Service general technical report no. PSW-43.
Available from: NTIS, Springfield, VA; PB81-133720.
Guderian, R.; Tingey, D. T.; Rabe, R. (1985) Effects of photochemical oxidants on plants. In: Guderian, R., ed.
Air pollution by photochemical oxidants. Berlin, Federal Republic of Germany:. Springer-Verlag;
pp. 127-333.
Hahn, J.; Crutzen, P. J. (1982) The role of fixed nitrogen in atmospheric photochemistry. Philos. Trans. R. Soc.
London B 296: 521-541.
Haines, E. B. (1979) Growth dynamics of cordgrass, Spartina alterniflora Loisel., on control and sewage sludge
fertilized plots in a Georgia salt marsh. Estuaries 2: 50-53.
Haines, T. A. (1987) Atlantic salmon resources in the northeastern United States and the potential effects of
acidification from atmospheric deposition. Water Air Soil Pollut. 35: 37-48.
Haines, S. G.; Cleveland, G. (1981) Seasonal variation in properties of five forest soils hi southwest Georgia.
Soil Sci. Soc. Am. J. 45: 139-143.
'- ' - • M t
Hanson, P. J.; Taylor, G. E., Jr. (1990) Modeling pollutant gas uptake by leaves: an approach based on
physicochemical properties. In: Dixon, R. K.; Meldahl, R.; Ruark, G.; lessen, N., eds. Forest growth:
process modeling of responses to environmental stress. Portland, OR: Timber Press; pp. 351-356.
Hanson, P. J.; Rott, K.; Taylor, G. E., Jr.; Gunderson, C. A.;. Lindberg, S. E.; Ross-Todd, B. M. (1989) NO2
deposition to elements representative of a forest landscape. Atmos. Environ. 23: 1783-1794.
Hanson, P. J.; Taylor, G. E., Jr.; Vose, J. (1991) Measurements of reactive nitrogen gas deposition to forest
landscape surfaces: biological and environmental controls. In: Johnson, D. W.; Lindberg, S. E., eds.
Atmospheric deposition and nutrient cycling in forest ecosystems. New York, NY: Springer-Verlag; in
press. • . • .. .
Hantschel, R.; Kaupenjohann, M.; Horn, R.; Zech, W. (1990) Water, nutrient and pollutant budgets in damaged
Norway spruce stands in NE-Bavaria (F.R.G.) and their changes after different fertilization treatments.
Water Air Soil Pollut. 49: 273-297. , • . '
Hao, W. M.; Wofsy, .S. C.; McElroy, M. B.; Beer, J. M.; Toqan,,M. A. (1987) Sources of atmospheric nitrous
. oxide from combustion. J. Geophys. Res. [Atmos.] 92: 3098-3104.
Harper, L. A.; Catchpoole, V. R.; Davis, R.; Weir, K. L. (1983) Ammonia volatilization: soil, plant, and
microclimate effects on diurnal and seasonal fluctuations. Agron. J. 75: 212-218.
Harper, L. A.; Sharpe, R. R.; Langdale, G. W.; Giddens, J. E. (1987) Nitrogen cycling in a wheat crop: soil,
plant, and aerial nitrogen transport. Agron. J. 79: 965-973.
Harrison, R. M.; Rapsomanikis, S.; Turnbull, A. (1989) Land-surface exchange in a chemically-reactive system;
. surface fluxes of HNO3, HC1, and NH3. Atmos. Environ. 23: 1795-1800.
Harriss, R. C.; Sebacher, D. L; Day, F. P., Jr. (1982) Methane flux in the Great Dismal Swamp. Nature
(London) 297: 673-674.
August 1991
10-255 DRAFT-DO NOT QUOTE OR CITE
-------
Harriss, R. C.; Gotham, E.; Sebacher, D. I.; Bartlett, K. B.; Flebbe, P. A. (1985) Methane flux from northern
peatlands. Nature (London) 315: 652-653.
Hauhs, M. (1989) Lange Bramke: an;ecosystem study of a forested watershed. In: Adriano, D. C.; Salmons, W.,
eds. Acidic precipitation. New York, NY: Springer-Verlag; pp. 275-305.
Hauhs, M.; Rost-Siebert, K.; Raben, G.; Paces, T.; Vigerust, B. (1989) Summary of European data. In:
Malanchuk, J. L.; Nilsson, L, eds. The role of nitrogen in the acidification of soils and surface waters.
Nordic Council of Ministers, Miljorapport 1989:10; pp. 5-1 - 5-37.
Healey, F. P. (1973) Inorganic nutrient uptake and deficiency in algae. Crit. Rev. Microbiol.: 69-113.
Hecky, R. E.; Kilham, P. (1988) Nutrient limitation of phytoplankton in freshwater and marine environments:
a review of recent evidence on trie effects of enrichment. Limnol. Oceanogr. 33: 796-822.
Hedin, L. O.; Likens, G. E.; Bormann, F. H. (1987) Decrease in precipitation acidity resulting from decreased
SO,,2" concentration. Nature (London) 325: 244-246.
Heil, G. W.; Bruggink, M. (1987) Competition for nutrients between Calluna vulgaris (L.) Hull and Molinia
caerulea (L.) Moench. Oecologia 73: 105-108.
Heil, G. W.; van Dam, D.; Heijne, B. (1987) Catch of atmospheric deposition in relation to vegetation structures
of heathland. In: Asman, W. A. H.; Diederen, H. S. M. A., eds. Ammonia and acidification:
proceedings of a symposium bf the European Association for the Science of Air Pollution (EURASAP);
April; Bilthoven, The Netherlands. European Association for the Science of Air Pollution; pp. 107-123.
Heil, G. W.; Werger, M. J. A.; de Mol, W.; van Dam, D.; Heijne, B. (1988) Capture of atmospheric
ammonium by grassland canopies. Science (Washington, DC) 239: 764-765.
Heilman, P. E,; Gessel, S. P. (1963) Nitrogen requirements and the biological cycling of nitrogen in Douglas-fir
stands in relationship to the effects of nitrogen fertilization. Plant Soil 18: 386-402.
Helvey, J. D.; Kunkle, S. H. (1986):Input-output budgets of selected nutrients on an experimental watershed near
Parsons, West Virginia. Broomall, PA: U. S. Department of Agriculture, Northeastern Forest Experiment
Station; USDA-Forest Service research paper NE-584.
Hemond, H. F. (1983) The nitrogen budget of Thoreau's Bog. Ecology 64: 99-109.
Henderson, G. S.; Harris, W. F. (1975) An ecosystem approach to characterization of the nitrogen cycle in a
deciduous forest watershed. In: Bernier, B.; Winget, C. H., eds. Forest soils and forest land
management: proceedings of the fourth North American forest soils conference; August 1973; Quebec,
PQ, Canada. Quebec, PQ, Canada: Les Presses de L'Universite Laval; pp. 179-183.
Hendrickson, O. Q. (1985) Variation in the C:N ratio of substrate mineralized during forest humus
decomposition. Soil Biol. Biochem. 17: 435-440.
Henriksen, A. (1988) Critical loads of nitrogen to surface waters. In: Nilsson, J.; Grennfelt, P., eds. Critical
loads for sulphur and nitrogen: report from a workshop; March; Skokloster, Sweden. Copenhagen,
Denmark: Nordic Council of Ministers; pp. 385-412.
Henriksen, A.; Brakke, D. F. (1988) Increasing contributions of nitrogen to the acidity of surface waters in
Norway. Water Air Soil Pollut. 42: 183-201.
August 1991
10-256 DRAFT-DO NOT QUOTE OR CITE
-------
Henriksen, A.; Lien, L.; Traaen, T. S.; Sevaldrad, I.'S.; Brakke, D. F. (1988) Lake acidification in Norway -
present and predicted .chemical status. Ambio 17: 259-266.
Hicks, B. B. (1989) Overview of deposition processes. In: Malanchuk, J. L.; Nilsson, J., eds. The role of
nitrogen in the acidification of soils and surface waters. Nordic Council of Ministers, Miljorapport
1989:10; pp. 3-1 - 3-21.
Hicks, B. B.; Matt, D. R. (1988) Combining biology, chemistry, and meteorology in modeling and measuring
dry deposition. J. Atmos. Chem. 6: 117-131.
Hicks, B. B.; Meyers, T. P. (1988) Measuring and modelling dry deposition in mountainous areas. In:
Unsworth, M. H.; Fowler, D., eds. Acid deposition at high elevation sites. Boston, MA: Kluwer
Academic Publishers; pp. 541-552.
Hicks, B. B.; Matt, D. R.; McMillen, R. T.; Womack, J. D.; Shelter, R. E. (1984) Eddy fluxes of nitrogen
oxides to a deciduous forest in complex terrain. In: Samson, P. J., ed. Transactions: the meteorology of
acid deposition, an APCA specialty conference; October 1983; Hartford, CT. Pittsburgh, PA: Air
Pollution Control Association; pp. 189-201.
Hicks, B. B.; Baldocchi, D. D.; Hosker, R. P., Jr.; Hutchison, B. A.; Matt, D. R.; McMillen, R. T.; .
Satterfield, L. C. (1985) On the use of monitored air concentrations to infer dry deposition (1985). Silver
Spring, MD: National Oceanic and Atmospheric Administration, Air Resources Laboratory; NOAA
technical memorandum no. ERL-ARL-141. Available from: NTIS, Springfield, VA; PB86-158409.
A Hi6ks,,B. B.; Baldocchi, D. D.; Meyers, T. P.; Hosker, R..P., Jr.; Matt, D. R. (1987) A preliminary multiple
resistance routine for deriving dry deposition velocities from measured quantities. Water Air Soil Pollut.
36: 311-330.
Hicks, B. B.; Matt, D. R.; McMillen, R. T. (1989) A micrometeorological investigation of surface exchange of
O3, SO2 and NO2: a case study. Boundary Layer Meteorol. 47: 321-336.
Hill, A. C. (1971) Vegetation: a sink for atmospheric pollutants. J. Air Pollut. Control Assoc. 21: 341-346.
Kingston, F. J.; Atkinson, R. J.; Posner, A- M.; Quirk, J. P. (1967) Specific adsorption of anions. Nature
(London) 215: 1459-1461.
Hoefken, ,K. D.; Gravenhorst, G. (1982) Deposition of atmospheric aerosol particles to beech- and spruce forest.
In: Georgii, H.-W.; Pankrath, J., eds. Deposition of atmospheric pollutants: proceedings of a
colloquium; November 1981; Oberursel/Taunus, Federal Republic of Germany. Dordrecht, The
Netherlands: D. ReideJ Publishing Company; pp. 191-194.
Hofstetter, R. H. (1983) Wetlands in the United States. In: Gore, Ai J. P., ed. Mires: swamp, bog, fen and
moor - regional studies. Amsterdam, The Netherlands: Elsevier Scientific Publishing Company;
pp. 201-244. (Ecosystems of the world 4B).
Hollis, C: A.; Smith, W. H.; Schmidt, R. A.; Pritchett, W. L. (1975) Soil and tissue nutrients, soil drainage,
fertilization and tree growth as related to fusiform rust incidence in slash pine. For. Sci. 21: 141-148.
• . .
Holmes, A. N.; Williams, W. D.; Wood, G. (1980) Relationships between forms of nitrogen and hydrological
characteristics in a small stream near Adelaide, South Australia. Aust. J; Mar. Freshwater Res.
31: 297-317.
Hosker, R. P., Jr.; Lindberg, S. E. (1982) Review: atmospheric deposition and plant assimilation of gases and
particles. Atmos. Environ. 16: 889-910.
August 1991
10-257 DRAFT-DO NOT QUOTE OR CITE
-------
Howarth, R. W. (1988) Nutrient limitation of net primary production in marine ecosystems. Annu. Rev. Ecol.
19: 89-110. ,
Howarth, R. W.; Cole, J. J. (1985) Molybdenum availability, nitrogen limitation, and phytoplankton growth in
natural waters. Science (Washington, DC) 229: 653-655.
Howarth, R. W.; Marino, R.; Lane, J.; Cole, J. J. (1988a) Nitrogen fixation in freshwater, estuarine, and
marine ecosystems. 1. Rates and importance. Limnol. Oceanogr. 33: 669-687.
Howarth, R. W.; Marino, R.; Cole, J. J. (1988b) Nitrogen fixation in freshwater, estuarine, and marine
ecosystems. 2. Biogeochemical controls. Limnol. Oceanogr. 33: 688-701.
Howes, B. L.; Dacey, J. W. H.; Goehringer, D. D. (1986) Factors controlling the growth form of Spartina
alterniflora: feedbacks between above-ground production, sediment oxidation, nitrogen and salinity.
J. Bool. 74: 881-898.
Hoyer, M. V.; Jones, J. R. (1983) Factors affecting the relation between phosphorus and chlorophyll a in
midwestem reservoirs. Can. J. Fish. Aquat. Sci. 40: 192-199.
Huebert, B. J. (1983) Measurements of the dry-deposition flux of nitric acid vapor to grasslands and forest. In:
Pruppacher, H. R.; Semonin, R.iG.; Slinn, W. G. N., eds. Precipitation scavenging, dry deposition, and
resuspension: volume 2, dry deposition and resuspension; proceedings of the fourth international,
conference; November-December 1982; Santa Monica, CA. New York, NY: Elsevier; pp. 785-794.
Huebert, B. J.; Robert, C. H. (1985) The dry deposition of nitric acid to grass. J. Geophys. Res. [Atmos.]
90: 2085-2090.
Huebert, B. J.; Luke, W. T.; Delany, A, C.; Brost, R. A. (1988) Measurements of concentrations and dry
surface fluxes of atmospheric nitrates in the presence of ammonia. J. Geophys. Res. [Atmos.]
93: 7127-7136. , . . ,
Husar, R. B. (1986) Emissions of sulfur dioxide and nitrogen oxides and trends for Eastern North America. In:
Acid deposition: long-term trends. Washington, DC: National Academy Press; pp. 48-92.
Hutchinson, G. E. (1973) Eutrophication: the scientific background of a contemporary practical problem. Am.
Sci. 61: 269-279.
Hutchinson, G. L.; Mosier, A. R. (1979) Nitrous oxide emissions from an irrigated,cornfield. Science ,
(Washington, DC) 205: 1125-1127.
Hutchinson, G. L.; Millington, R. J.; Peters, D. B. (1972) Atmospheric ammonia: absorption by plant leaves.
Science (Washington, DC) 175: 771-772.
Ingestad, T. (1980) Nutrition and growth of birch and gray alder seedlings in low conductivity solutions and at
varied relative rates of nutrient addition. Physiol. Plant. 52: 454-466.
James, B. R.; Riha, S. J. (1989) Aluminum leaching by mineral acids in forest soils: I. nitric-sulfuric acid
differences. Soil Sci. Soc. Am. J. 53: 259-264.
Jarvis, P. G. (1971) The estimation of resistances to carbon dioxide transfer. In: Sestak, Z.; Catsky, J.; Jarvis,
P. G., eds. Plant photosynthetic production: manual of methods. The Hague, The Netherlands:
Dr W. Junk N.V. Publishers; pp. 566, 608-609. -
August 1991
10-258
DRAFT-DO NOT QUOTE OR CITE
-------
Jarvis, P. G.; Catsky, J.; Eckardt, F. E.; Koch, W.; Koller, D. (1971) General principles of gasometric methods
and the main aspects of installation design. In: Sestak, Z.; Catsky, J.; Jarvis, P. G., eds. Plant
photosynthetic production: manual of methods. The Hague, The Netherlands: Dr W. Junk N. V.
• Publishers; pp. 49-110.
Jaworski, N. A. (1981) Sources of nutrients and the scale of eutrophication problems in estuaries. In: Neilson,
B. J.; Cronin, L. E., eds. Estuaries and nutrients. Clifton, NJ: Humana Press; pp. 83-110.
Jefferies, R. L.; Perkins, N. (1977) The effects on the vegetation of the additions of inorganic nutrients to salt
marsh soils at Stiffkey, Norfolk. J. Ecol. 65:^867-882.
Jeffries, D. S. (1990) Snowpack storage of pollutants, release during melting, and impact on receiving waters. In:
Norton, S. A.; Lindberg, S. E;; Page, A. L., eds. Acidic precipitation: v. 4, soils, aquatic processes,
and lake acidification. New York, NY: Springer-Verlag; pp. 107-132.
Jenkins, M. C.; Kemp, W. M. (1984) The coupling of nitrification and denitrification in two estuarine
sediments. Limnol. Oceanogr. 29: 609-619. "
' ' • '.;•'.
Jenkinson, D. S. (1970) The accumulation of organic matter in soil left uncultivated. Comm. Bur. Soil Sci.
Rothamstad Exp. Sta. Rep. for 1970.
Jensen,' A. (1985) The effect of cattle and sheep grazing on salt-marsh vegetation at Skallingen, Denmark.
Vegetatio60: 37-48. .'••-'.
Johannessen, M.; Henriksen, A. (1978) Chemistry of snow meltwater: changes in concentration during melting.
Water Resour. Res. 14: 615-619.
Johannessen, M.; Skartveit, A.; Wright, R. F. (1980) Streamwater chemistry before, during and after snowmelt.
In: Drablos, D.; Tollan, A., eds. Ecological impact of acid precipitation: proceedings of an international
conference; March; Sandefjord, Norway. Oslo, Norway: SNSF project; pp. 224-225.
Johansson, C. (1987) Pine forest: a negligible sink for atmospheric NOX in rural Sweden. Tellus Ser. B
39: 426-438. , '
Johansson; C.; Granat, L. (1986) An experimental study of the dry deposition Of gaseous nitric acid to snow..
Atmos. Environ. 20: 1165-1170.
John, W.; Wall, S. M.; Ondo, J. L. (1985) Dry acid deposition on materials and vegetation: concentrations in
ambient air. Berkeley, CA: California Department of Health Services, Air and Industrial Hygiene
Laboratory; report no. CA/DOH/AIHL/SP-34.
Johnson, A. H. (1983) Red spruce decline in the northeastern U.S.: hypotheses regarding the role of acid ram.
J. Air Pollut. Control Assoc. 33: 1049-1054.
Johnson, D. W.; Edwards, N. T. (1979) The effects of stem girdling on biogeochemical cycles within a mixed
deciduous forest in eastern Tennessee: II. soil nitrogen mineralization and nitrification rates. Oecologia
40: 259-271. : '
Johnson, D. W.; Lindberg, S. E., eds. (n.d.) Atmospheric deposition and nutrient cycling in forest ecosystems.
New York, NY: Springer-Verlag; in press. • *
Johnson, D. W.; Todd, D. E. (1987) Nutrient export by leaching and whole-tree harvesting in a loblolly pine
and mixed oak forest. Plant Soil 102: 99-109.
August 1991
10-259 DRAFT-DO NOT QUOTE OR CITE
-------
Johnson, D. W.; Todd, D. E. (1988) Nitrogen fertilization of young yellow poplar and loblolly pine plantations
at differing frequencies. Soil Sci. Soc. Am. J. 52: 1468-1477.
Johnson, D. W.; van Miegroet, H.; Lindberg, S. E.; Harrison, R. B.; Todd, D. E. (n.d.) Nutrient cycling in red
spruce forests of the Great Smoky Mountains. Can. J. For. Res.: in press.
Johnson, N. M.; Likens, G. E.; Bormann, F. H.; Fisher, D. W.; Pierce, R. S. (1969) A working model for the
variation in stream water chemistry at the Hubbard Brook Experimental Forest, New Hampshire. Water
Resour. Res. 5: 1353-1363.
Johnson, D. W.; Cole, D. W.; Gessel, S. P.; Singer, M. J.; Minden, R. V. (1977) Carbonic acid leaching in a
tropical, temperate, subalpine and northern forest soil. Arct. Alp. Res. 9: 329-343.
Johnson, N. M.; Driscoll, C. T.; Eaton, J. S.; Likens, G. E.; McDowell, W. H. (1981) "Acid rain", dissolved
aluminum and chemical weathering at the Hubbard Brook Experimental Forest, New Hampshire.
Geochim. Cosmochim. Acta 45: 1421-1437.
Johnson, D. W.; Richter, D. D.; Lovett, G. M.; Lindberg, S. E. (1985) The effects of atmospheric deposition
on potassium, calcium, and magnesium cycling in two deciduous forests. Can. J. For. Res. 15: 773-782.
Jones, H. C.; Noggle, J. C.; Young, R. C.; Kelly, J. M.; Olem, H.; Ruane, R. J.; Pasch, R. W.; Hyfantis,
G. J.; Parkhurst, W. J. (1983) Investigations of the cause of fishkills in fish-rearing facilities in Raven
Fork watershed. Chattanooga, TN: Tennessee Valley Authority, Office of Natural Resources; report no.
TVA/ONR/WR-83-9.
Jordan, T. E.; Correll, D. L.; Whigham, D. F. (1983) Nutrient flux in the Rhode River: tidal exchange of
nutrients by brackish marshes. Estuarine Coastal Shelf Sci. 17: 651-667.
Joslin, J. D.; Wolfe, M. H. (1988) Responses of red spruce seedlings to changes to soil aluminum in six
amended forest soil horizons. Can. J. For. Res. 18: 1614-1623.
Joslin, J. D.; Mays, P. A.; Wolfe, M.iH.; Kelly, J. M.; Garber, R. W.; Brewer, P. F. (1987) Chemistry of
tension lysimeter water and lateral flow in spruce and hardwood stands. J. Environ. Qual. 16: 152-160.
Judeikis, H. S.; Wren, A. G. (1978) Laboratory measurements of NO and NO2 depositions onto soil and cement
surfaces. Atmos. Environ. 12: 2315-2319.
Jurgensen, M. F.; Harvey, A. E.; Larsen, M. J. (1981) Effects of prescribed fire on soil nitrogen levels in a
cutover Douglas-fir/western larch forest. Ogden, UT: U. S. Department of Agriculture Forest Service,
Intermt. Forest and Range Experiment Station; research paper INT-275.
Kadlec, J. A. (1986) Input-output nutrient budgets for small diked marshes. Can. J. Fish. Aquat. Sci.
43: 2009-2016.
Kahl, J. S.; Norton, S. A.; Cronan, C. A.; Fernandez, I. J.; Bacon, L. C.; Haines, T. A. (1991) Maine. In:
Charles, D. F., ed. Acidic deposition and aquatic ecosystems: regional case studies. New York, NY:
Springer Verlag.
Kanciruk, P.; Eilers, J. M.; McCord, R. A.; Landers, D. H.; Brakke, D. F.; Linthurst, R. A. (1986)
Characteristics of lakes in the eastern United States. Volume III: data compendium of site characteristics
and chemical variables. Washington, DC: U. S. Environmental Protection Agency, Office of Acid
Deposition, Environmental Monitoring, and Quality Assurance; EPA report no. EPA-600/4-86-007c.
Available from: NTIS, Springfield, VA; PB87-110409/REB.
August 1991
10-260
DRAFT-DO NOT QUOTE OR CITE
-------
Katz, B. G.; Bricker, O. P.; Kennedy, M. M. (1985) Geochemical mass-balance relationships for selected ions in
precipitation and stream water, Catoctin Mountains, Maryland. Am. J. Sci. 285: 931-962.
Kaufmann, P. R.; Herlihy, A. T.; Elwood, J. W.; Mitch, M. E.; Overton, W. S.; Sale, M. J.; Messer, J. J.;
Cougan, K. A.; Peck, D. V.; Reckhow, K. H.; Kinney, A. J.; Christie, S. J.; Brown, D. D.; Hagley,
C. A.; Jager, H. I. (1988) Chemical characteristics of streams in the mid-Atlantic and southeastern
United States (national stream survey - phase I). Volume I: Population descriptions and physico-chemical
relationships. Washington, DC: U. S. Environmental Protection Agency, Office of Acid Deposition,
Environmental Monitoring and Quality Assurance; EPA report no. EPA-600/3-88-021a.
Kaufmann, P. R.; Herlihy, A. T.; Mitch, M. E.; Messer, J. J.; Overton, W. S. (1991) Stream chemistry in the
eastern United States: 1. synoptic survey design, acid-base status, and regional patterns. Water Resour.
Res. 27: 611-627.
Kaushik, N. K.; Robinson, J. B.; Sain, P.; Whitely, H. R.; Stammers, W. N. (1975) A quantitative study of
nitrogen loss from water of a small spring-fed stream. Water Pollut. Res. J. Can. 10: 110-117.
Keddy, P. A.; Wisheu, I. C. (1989) Ecology, biogeography, and conservation of coastal plain plants: some
general principles from the study of Nova Scotian wetlands. Rhodora 91: 72-94.
Keeney, D. R.; Chen, R. L.; Graetz, D. A. (1971) Importance of denitrification and nitrate reduction in
sediments to the nitrogen budgets of lakes. Nature (London) 233: 66-67.
Kelly, J. M. (1988) Annual elemental input/output estimates for two forested watersheds in eastern Tennessee.
J. Environ. Qual. 17: 463-468.
Kelly, J. M.; Henderson, G. S. (1978) Effects of nitrogen and phosphorus additions on deciduous litter
decomposition. Soil Sci. Soc. Am. J. 42: 972-976.
Kelly, J. M.; Meagher, J. F. (1986) Nitrogen input/output relationships for three forested watersheds in eastern
Tennessee. In: Correll, D. L., ed. Watershed research perspectives. Washington, DC: Smithsonian
Institution Press; pp. 360-391.
Kelly, C. A.; Rudd, J. W. M.; Cook, R. B.; Schindler, D. W. (1982) The potential importance of bacterial
processes in regulating rate of lake acidification. Limnol. Oceanogr. 27: 868-882.
Kelly, T. J.; Tanner, R. L.; Newman, L.; Galvin, P. J.; Kadlecek, J. A. (1984) Trace gas and aerosol
measurements at a remote site in the northeast U.S. Atmos. Environ. 18: 2565-2576.
Kelly, T. J.; McLaren, S. E.; Kadlecek, J. A. (1989) Seasonal variations in atmospheric SOX and NOy species in
the Adirondacks. Atmos. Environ. 23: 1315-1332.
Kelly, C. A.; Rudd, J. W. M.; Schindler, D. W. (1990) Acidification by nitric acid - future considerations.
Water Air Soil Pollut. 50: 49-61.
Kenk, G.; Fischer, H. (1988) Evidence from nitrogen fertilisation in the forests of Germany. Environ. Pollut.
54: 199-218.
Kilham, P.; Hecky, R. E. (1988) Comparative ecology of marine and freshwater phytoplankton. Limnol.
Oceanogr. 33: 776-795.
King, D.; Nedwell, D. B. (1985) The influence of nitrate concentration upon the end-products of nitrate
dissimilation by bacteria in anaerobic salt marsh sediment. FEMS Microbiol. Ecol. 31: 23-28.
August 1991
10-261
DRAFT-DO NOT QUOTE OR CITE
-------
Kisser-Priesack, G. M.; Scheunert, I.; Gnatz, G.; Ziegler, H. (1987) Uptake of 1SNO2 and 1SNO by plant
cuticles. Naturwissenschaften 74: 550-551.
Klein, R. M.; Perkins, T. D. (1987) Cascades of causes and effects of forest decline. Ambio 16: 86-93.
Klein, T. M.; Kreitinger, J. P.; Alexander, M. (1983) Nitrate formation in acid forest soils from the
Adirondacks. Soil Sci. Soc. Am. J. 47: 506-508.
Klemedtsson, L.; Svensson, B. H. (1988) Effects of acid deposition on denitrification and N2O-emission from
forest soils. In: Nilsson, J.; Grennfelt, P., eds. Critical loads for sulphur and nitrogen: report from a
workshop; March; Skokloster, Sweden. Copenhagen, Denmark: Nordic Council of Ministers;
pp. 343-362.
Klepper, L. (1979) Nitric oxide (NO) and nitrogen dioxide (NOj) emissions from herbicide-treated soybean
plants. Atmos. Environ. 13: 537-542.
Klubek, B.; Eberhardt, P. J.; Skujins, J. (1978) Ammonia volatilization from Great Basin desert soils. In: West,
N. E.j Skujins, J., eds. Nitrogen in desert ecosystems. Stroudsburg, PA: Dowden, Hutchinson & Ross,
Inc.; pp. 107-131. (US/IBP synthetic series 9).
Knowles, R. (1982) Denitrification. Microbiol. Rev. 46: 43-70.
Koerselman, W.; Bakker, S. A.; Blom; M. (1990) Nitrogen, phosphorus and potassium budgets for two small
fens surrounded by heavily fertilized pastures. J. Ecol. 78: 428-442.
Kolb, W.; Martin, P. (1988) Influence of nitrogen on the number of N2-fixing and total bacteria in the
rhizosphere. Soil Biol. Biochem. 20: 221-225.
Kowalenko, C. G.; Ivarson, K. C.; Cameron, D. R. (1978) Effect of moisture content, temperature and nitrogen
fertilization on carbon dioxide evolution from field soils. Soil Biol. Biochem. 10: 417-423.
Kozlowski, T. T. (1980) Impacts of air pollution on forest ecosystems. Bioscience 30: 88-93.
Kramer, J. R.; Andren, A. W.; Smith, R. A.; Johnson, A. H.; Alexander, R. B.; Oehlert, G. (1986) Streams
and lakes. In: Acid deposition: long-term trends. Washington, DC: National Academy Press;
pp. 231-299.
Kramm, G. (1989) A numerical method for determining the dry deposition of atmospheric trace gases. Boundary
Layer Meteorol. 48: 157-175.
Krause, G. H. M. (1988) Impact of air pollutants on above-ground plant parts of forest trees. In: Mathy, P., ed.
Air pollution and ecosystems: proceedings of an international symposium; May 1987; Grenoble, France.
Boston, MA: D. Reidel Publishing Company; pp. 168-197.
Kuenzler, E. J.; Craig, N. J. (1986) Land use and nutrient yields of the Chowan River watershed. In: Correll,
D. L., ed. Watershed research perspectives. Washington, DC: Smithsonian Institution Press; pp. 77-107.
Lamb, D. (1980) Soil nitrogen mineralisation in a secondary rainforest succession. Oecologia 47: 257-263.
August 1991
10-262
DRAFT-DO NOT QUOTE OR CITE
-------
Landers, D. H.; Eilers, J. M.; Brakke, D. F.; Overton, W. S.; Kellar, P. E.; Silverstein, M. E.; Schbnbrod,
R. D.; Crowe, R. E.; Linthurst, R. A.; Omernik, J. M.; Teague, S. A.; Meier, E. P. (1987) Western
lake survey phase I: characteristics of lakes in the western United States, volume I: population
descriptions and physico-chemical relationships. Washington, DC: U. S. Environmental Protection
Agency, Office of Acid Deposition, Environmental Monitoring and Quality Assurance; EPA report no.
EPA/600/3-86-054A. Available from: NTIS, Springfield, VA; PB88-146824/REB.
Landsberg, J. J. (1986) Experimental approaches to the study of the effects of nutrients and water on carbon
assimilation by trees. Tree Physiol. 2: 427-444.
Larsson, U.; Elmgren, R.; Wulff, F. (1985) Eutrophication and the Baltic Sea: causes and consequences. Ambio
14: 9-14.
Law, R. M.; Mansfield, T. A. (1982) Oxides of nitrogen and the greenhouse atmosphere. In: Unsworth, M. H.;
Ormrod, D. P., eds. Effects of gaseous air pollution in agriculture and horticulture. London, United
Kingdom: Butterworth Scientific; pp. 93-112.
LaZerte, B. D.; Dillon, P. J. (1984) Relative importance of anthropogenic versus natural sources of acidity in
lakes and streams of central Ontario. Can. J. Fish. Aquat. Sci. 41: 1664-1677.
Lee, J. A.; Press, M. C.; Woodin, S. J. (1986) Effects of NO2 on aquatic ecosystems. In: Environment and
quality of life: study on the need for an NO2 long-term limit value for the protection of terrestrial and
' aquatic ecosystems. Luxembourg: Commission of the European Communities; pp. 99-119.
Lefohn, A. S.; Tingey, D. T. (1984) The co-occurrence of potentially phytotoxic concentrations of various
gaseous air pollutants. Atmos. Environ. 1.8: 2521-2526.
Lemon, E.; van Houtte, R. (1980) Ammonia exchange at the land surface. Agron. J. 72: 876-883.
Lendzian, K. J.; Kerstiens, G. (1988) Interactions between plant cuticles and gaseous air pollutants. Aspects
Appl. Biol. 17: 97-104.
Leuven, R. S. E. W.; Wolfs, W. J. (1988) Effects of water acidification on the decomposition ofJuncus
bulbosus L. Aquat. Bot... 31: 57-81.
Levy, H., II; Moxim, W. J. (1987) Fate of US and Canadian combustion nitrogen emissions. Nature (London)
328: 414-416.
Likens, G. E., ed. (1972) Nutrients and eutrophication: the limiting-nutrient controversy: proceedings of the
symposium; Feb. 1971; East Lansing, MI. Lawrence, KS: American Society of Limnology and
Oceanography, Inc. (Special symposia volume I).
Likens, G. E., ed. (1985) An ecosystem approach to aquatic ecology: Mirror Lake and its environment. New
York, NY: Springer-Verlag.
Likens, G. E.; Bormann, F. H.; Johnson, N. M.; Fisher, D. W.; Pierce, R. S. (1970) Effects of forest cutting
and herbicide treatment on nutrient budgets in the Hubbard Brook watershed-ecosystem. Ecol. Monogr.
40:23-47.
Likens, G. E.; Bormann, F. H.; Pierce, R. S.; Eaton, J. S.; Johnson, N. M. (1977) Biogeochemistry of a
forested ecosystem. New York, NY: Springer-Verlag, Inc.
Likens, G. E.; Bormann, F. H.; Pierce, R. S.; Eaton, J. S.; Munn, R. E. (1984) Long-term trends in
precipitation chemistry at Hubbard Brook, New Hampshire. Atmos. Environ. 18: 2641-2647.
August 1991
10-263
DRAFT-DO NOT QUOTE OR CITE
-------
Liljelund, L.-E.; Torstensson, P. (1988) Critical load of nitrogen with regards to effects on plant composition.
In: Nilsson, J.; Grennfelt, P., eds. Critical loads for sulphur and nitrogen: report from a workshop;
March; Skokloster, Sweden. Copenhagen, Denmark: Nordic Council of Ministers; pp. 363-373.
Lindau, C. W.; De Laune, R. D.; Jones, G. L. (1988) Fate of added nitrate and ammonium-nitrogen entering a
Louisiana gulf coast swamp forest. J. Water Pollut. Control Fed. 60: 386-390.
Lindberg, S. E.; Lovett, G. M. (1985) Field measurements of particle dry deposition rates to foliage and inert
surfaces in a forest canopy. Environ. Sci. Technol. 19: 238-244.
Lindberg, S. E.; Lovett, G. M.; Richter, D. D.; Johnson, D. W. (1986) Atmospheric deposition and canopy
interactions of major ions in a forest. Science (Washington, DC) 231: 141-145.
Lindberg, S. E.; Lovett, G. M.; Meiwes, K.-J. (1987) Deposition and forest canopy interactions of airborne
nitrate. In: Hutchinson, T. C.; Meema, K. M., eds. Effects of atmospheric pollutants on forests,
wetlands and agricultural ecosystems. Berlin, Federal Republic of Germany: Springer-Verlag;
pp. 117-130. (NATO as 1 series: v. G16).
Lindsay, W. L. (1979) Chemical equilibria in soils. New York, NY: John Wiley & Sons, Inc.
Linsey, G. A.; Schindler, D. W.; Stainton, M. P. (1987) Atmospheric deposition of nutrients and major ions at
the Experimental Lakes Area in northwestern Ontario, 1970 to 1982. Can. J. Fish. Aquat. Sci. 44
(suppl. 1): 206-214.
Linthurst, R. A.; Landers, D. H.; Eilers, J. M.; Brakke, D. F.; Overton, W. S.; Meier, E. P.; Crowe, R. E.
(1986) Characteristics of lakes in the eastern United States. Volume I. Population descriptions and
physico-chemical relationships. Washington, DC: U. S. Environmental Protection Agency, Office of Acid
Deposition, Environmental Monitoring, and Quality Assurance; EPA report no. EPA-600/4-86-007a.
Available from: NTIS, Springfield, VA; PB87-110383.
Lockyer, D. R.; Whitehead, D. C. (1986) The uptake of gaseous ammonia by the leaves of Italian ryegrass.
J. Exp. Bot. 37: 919-927.
Lodhi, M. A. K. (1978) Comparative inhibition of nitrifiers and nitrification in a forest community as a result of
the allelopathic nature of various tree species. Am. J. Bot. 65: 1135-1137.
Loftis, J. S.; Ward, R. C.; Phillips, R. D.; Taylor, C. H. (1989) An evaluation of trend detection techniques for
use in water quality monitoring programs. Research Triangle Park, NC: U. S. Environmental Protection
Agency; EPA report no. EPA-600/3-89-037.
Logofet, D. O.; Alexandrov, G. A. (1984) Modelling of matter cycle in a mesotrophic bog ecosystem:
n. dynamic model and ecological succession. Ecol. Modell. 21: 259-276.
Loranger, T. J.; Brakke, D. F. (1988) The extent of snowpack influence on water chemistry in a North Cascades
lake. Water Resour. Res. 24: 723-726.
Loranger, T. J.; Brakke, D. F.; Bonoff, M. B.; Gall, B. F. (1986) Temporal variability of lake waters in the
North Cascades Mountains (Washington, U.S.A.). Water Air Soil Pollut. 31: 123-129.
Lovett, G. (1991) Atmospheric deposition and canopy interactions of nitrogen. In: Johnson, D. W.; Lindberg,
S. E., eds. Atmospheric deposition and nutrient cycling in forest ecosystems. New York, NY:
Springer-Verlag; in press.
August 1991
10-264 DRAFT-DO NOT QUOTE OR CITE
-------
Lovett, G. M.; Lindberg, S. E. (1984) Dry deposition and canopy exchange in a mixed oak forest as determined
by analysis of throughfall. J. Appl. Ecol. 21: 1013-1027.
Lovett, G. M.; Lindberg, S. E. (1986) Dry deposition of nitrate to a deciduous forest. Biogeochemistry
2: 137-148.
Lovett, G. M.; Reiners, W. A.; Olson, R. K. (1982) Cloud droplet deposition in subalpine balsam fir forests:
hydrological and chemical inputs. Science (Washington, DC) 218: 1303-1304.
Lbwranee, R. R.; Leonard, R. A. (1988) Streamflow nutrient dynamics on coastal plain watersheds. J. Environ.
Qual. 17: 734-740.
Lowrance, R. R.; Leonard, R. A.; Asmussen, L. E.; Todd, R. L. (1985) Nutrient budgets for agricultural
watersheds in the southeastern coastal plain. Ecology 66: 287-296.
Loye-Pilot, M. D.; Martin, J. M.; Morelli, J. (1990) Atmospheric input of inorganic nitrogen to the Western
Mediterranean. Biogeochemistry 9: 117-134.
Lynch, J. A.; Hanna, C. M.; Corbett, E. S. (1986) Predicting pH, alkalinity, and total acidity in stream water
during episodic events. Water Resour. Res. 22: 905-912.
Manny, B. A.; Owens, R. W. (1983) Additions of nutrients and major ions by the atmosphere and tributaries to
nearshore waters of northwestern Lake Huron. J. Great Lakes Res. 9: 403-420.
Margolis, H. A.; Waring, R. H. (1986) Carbon and nitrogen allocation patterns of Douglas-fir seedlings
fertilized with nitrogen in autumn. II. Field performance. Can. J. For. Res. 16: 903-909.
Marinucci, A. C.; Hobbie, J. E.; Helfrich, J. V. K. (1983) Effect of litter nitrogen on decomposition and
microbial biomass in Spartina alterniflora. Microb. Ecol. 9: 27-40.
Marshall, J. D.; Cadle, S. H. (1989) Evidence for trans-cuticular uptake of HNO3 vapor by foliage of eastern
white pine (Pinus strobus L.). Environ. Pollut. 60: 15-28.
Matson, P. A.; Harriss, R. C. (1988) Prospects for aircraft-based gas exchange measurements in ecosystem
studies. Ecology 69: 1318-1325.
Matson, P. A.; Vitousek, P. M. (1990) Ecosystem approach to a global nitrous oxide budget. Bioscience
40: 667-672.
Matzner, E. (1983) Balances of element fluxes within different ecosystems impacted by acid rain. In: Ulrich, B.;
Pankrath, J., eds. Effects of accumulation of air pollutants in forest ecosystems: proceedings of a
workshop; May 1982; Goettingen, Federal Republic of Germany. Dordrecht, The Netherlands: D. Reidel
Publishing Company; pp. 147-155.
Matzner, E.; Khanna, P. K.; Meiwes, K. J.; Ulrich, B. (1983) Effects of fertilization on the fluxes of chemical
elements through different forest ecosystems. Plant Soil 74: 343-358.
Mazzarino, M. J.; Heinrichs, H.; Foelster, H. (1983) Holocene versus accelerated actual proton consumption in
German forest soils. In: Ulrich, B.; Pankrath, J., eds. Effects of accumulation of air pollutants in forest
ecosystems: proceedings of a workshop; May 1982; Goettingen, Federal Republic of Germany.
Dordrecht, The Netherlands: D. Reidel Publishing Company; pp. 113-123.
McAvoy, D. C. (1989) Episodic response of aluminum chemistry in an acid-sensitive Massachusetts catchment.
Water Resour. Res, 25: 233-240.
August 1991
10-265
DRAFT-DO NOT QUOTE OR CITE
-------
McCarthy, J. J. (1981) The kinetics of nutrient utilization. Can., Bull. Fish. Aquat. Sci. 210: 211-233..
McCauley, E.; Downing, J. A.; Watson, S. (1989) Sigmoid relationships between nutrients and chlorophyll
among lakes. Can. J. Fish. Aquat. Sci. 46: 1171-1175.
McComb, A. J.; Atkins, R. P.; Birch, P. B.; Gordon, D. M.; Lukatelich, R. J. (1981) Eutrophication in the
Peel-Harvey estuarine system, Western Australia. In: Neilson, B. J.; Cronin, L. E., eds. Estuaries and
nutrients. Clifton, NJ: Humana Press; pp. 323-342.
McHugh, J. L. (1976) Estuarine fisheries: are they doomed? In: Wiley, M. L., ed. Estuarine processes: v. 1.
New York, NY: Academic Press; pp. 15-27.
McLaughlin, S. B. (1985) Effects of air pollution on forests: a critical review. J. Air Pollut. Control Assoc.
35: 512-534.
McMahon, T. A.; Denison, P. J. (1979) Empirical atmospheric deposition parameters—a survey. Atmos.
Environ. 13: 571-585.
McMillen, R. T. (1988) An eddy correlation technique with extended applicability to non-simple terrain.
Boundary Layer Meteorol. 43: 231-245.
McNulty, S. G.; Aber, J. D.; McLellan, T. M.; Katt, S. M. (1990) Nitrogen cycling in high elevation forests of
the northeastern US in relation to nitrogen deposition. Ambio 19: 38-40.
Mead, D. J.; Pritchett, W. L. (1975) Fertilizer movement in a slash pine ecosystem: II. N distribution after two
growing seasons. Plant Soil 43: 467-478.
Meams, A. J.; Haines, E.; Kleppel, G. S.; McGrath, R. A.; McLaughlin, J. J. A.; Segar, D. A.; Sharp, J. H.;
Walsh, J. J.; Word, J. Q.; Young, D. K.; Young, M. W. (1982) Effects of nutrients and carbon loadings
on communities and ecosystems. In: Mayer, G. F., ed. Ecological stress and the New York Bight:
science and management: proceedings of a symposium on the ecological effects of environmental, stress;
June 1979; New York, NY. Columbia, SC: Estuarine Research Federation; pp. 53-65.
Melack, J. M.; Stoddard, J. L. (1991) Sierra Nevada. In: Charles, D. F., ed. Acidic deposition and aquatic
ecosystems: regional case studies. New York, NY: Springer Verlag; In press.
Melack, J. M.; Kilham, P.; Fisher, T. R. (1982) Responses of phytoplankton to experimental fertilization with
ammonium and phosphate in an African soda lake. Oecologia 52: 321-326.
Melack, J. M.; Cooper, S. D.; Holmes, R. W.; Sickman, J. O.; Kratz, K.; Hopkins, P.; Hardenbergh, H.;
Thime, M.; Meeker, L. (1987) Chemical and biological survey of lakes and streams located in the
Emerals Lake watershed, Sequoia National Park. Sacramento, CA: California Air Resources Board;
report no. A3-096-32.
Melillo, J. M.; Aber, J. D.; Steudler, P. A.; Schimel, J. P. (1983) Denitrification potentials in a successional
sequence of northern hardwood forest stands. In: Hallberg, R., ed. Environmental biogeochemistry:
proceedings of the 5th international symposium on environmental biogeochemistry; June 1981;
Stockholm, Sweden. Ecol. Bull. 35: 217-228.
Melillo, J. M.; Naiman, R. J.; Aber, J. D.; Linkins, A. E. (1984) Factors controlling mass loss and nitrogen
dynamics of plant litter decaying in northern streams. Bull. Mar. Sci. 35: 341-356.
August 1991
10-266
DRAFT-DO NOT QUOTE OR CITE
-------
Melillo, J. M.; Steudler, P. A.!; Aber, J. D.; Bowden, R. D. (1989) Atmospheric deposition and nutrient
cycling. In: Andreae, M. O.; Schimel, D. S., eds. Exchange of trace gases between terrestrial ecosystems
and the atmosphere: report of the Dahlem workshop; February; Berlin, Federal Republic of Germany.
New York, NY: John Wiley & Sons, Ltd.; pp. 263-280.
Melin, J.; Nommik, H.; Lohm, U.; Flower-Ellis, J. (1983) Fertilizer nitrogen budget in a Scots pine ecosystem
attained by using root-isolated plots and 15N tracer technique. Plant Soil 74: 249-263.
Menge, J. A.; Grand, L., F.; Haines, L. W. (1977) The effect of fertilization on growth and mycorrhizae
numbers in 11-year-old loblolly pine plantations. For. Sci. 23: 37-44.
Messer, J. J.; Ho, J.; Grenney, W. J. (1984) Ionic strength correction for extent of ammonia ionization in
freshwater. Can. J. Fish. Aquat. Sci. 41: 811-815.
Messer, J. J.; Ariss, C. W.; Baker, J. R.; Drouse, S. K.; Eshleman, K. N.; Kinney, A. J.; Overton, W. S.;
Sale, M. J.; Schonbrod, R. D. (1988) Stream chemistry in the southern Blue Ridge: feasibility of a
regional synoptic sampling approach. Water Resour. Bull. 24: 821-829.
Meyer, J. L. (1979) The role of sediments and bryophytes in phosphorus dynamics in a headwater stream
ecosystem. Limnol. Oceanogr. 24: 365-375.
Meyer, J. L.; Likens, G. E. (1979) Transport and transformation of phosphorus in a forest stream ecosystem.
Ecology 60: 1255-1269.
Meyers, T. P.; Hicks, B. B. (1988) Dry deposition of O3, SO2, and HNO3 to different vegetation in the same
exposure environment. Environ. Pollut. 53: 13-25.
Meyers, T. P.; Huebert, B. J.; Hicks, B. B. (1989) HNO3 deposition to a deciduous forest. Boundary Layer
Meteorol. 49: 395-410.
Miller, P. L. (1973) Oxidant-induced community change in a mixed conifer forest. In: Naegele, J. A., ed. Air
pollution damage to vegetation. Washington, DC: American Chemical Society; pp. 101-117. (Advances
in chemistry series: no. 122).
Miller, H. G. (1981) Forest fertilization: some guiding concepts. Forestry 54: 157-167.
Miller, H. G.; Miller, J. D. (1988) Response to heavy nitrogen applications in fertilizer experiments in British
forests. Environ. Pollut. 54: 219-231.
Miller, H. G.; Miller, J. D.; Pauline, O. J. L. (1976) Effect of nitrogen supply on nutrient uptake in Corsican
pine. J. Appl. Ecol. 13: 955-966.
Mitsch, W. J.; Gosselink, J. G. (1986) Wetlands. New York, NY: Van Nostrand Reinhold Company.
Mollitor, A. V.; Raynal, D. J. (1982) Atmospheric deposition and ionic input in Adirondack forests. J. Air
Pollut. Control Assoc. 33: 1032-1036.
Molot, L. A.; Dillon, P. J.; LaZerte, B. D. (1989) Factors affecting alkalinity concentrations of streamwater
during snowmelt in central Ontario. Can. J. Fish. Aquat. Sci. 46: 1658-1666.
Montagnini, F.; Haines, B. L.; Swank, W. T.; Waide, J. B. (1989) Nitrification in undisturbed mixed
hardwoods and manipulated forests in the southern Appalachian Mountains of North Carolina, USA. Can.
J. For. Res. 19: 1226-1234.
August 1991
10-267 DRAFT-DO NOT QUOTE OR CITE
-------
Moore, B., HI; Bolin, B. (1986) The oceans, carbon dioxide, and global climate change. Oceanus 29(4): 9-15.
Moore, D. R. J.; Keddy, P. A. (1989) The relationship between species richness and standing crop in wetlands:
the importance of scale. Vegetatio 79: 99-106.
Moore, I. D.; Nuckols, J. R. (1984) Relationship between atmospheric nitrogen deposition and the stream
nitrogen profile. J. Hydrol. 74: 81-103.
Moore, D. R. J.; Keddy, P. A.; Gaudet, C. L.; Wisheu, I. C. (1989) Conservation of wetlands: do infertile
wetlands deserve a higher priority? Biol. Conserv. 47: 203-217.
Moorhead, D. L.; Reynolds, J. F.; Whitford, W. G. (1986) A conceptual model for primary productivity,
decomposition and nitrogen cycling in the Chihuahuan creosotebush desert. Tree Physiol. 2: 215-222.
Morgan, M. D.; Good, R. E. (1988) Stream chemistry in the New Jersey Pinelands: the influence of
precipitation and watershed disturbance. Water Resour. Res. 24: 1091-1100.
Morris, J. T. (1982) A model of growth responses by Spartina alterniflora to nitrogen limitation. J. Ecol.
70: 25-42.
Morris, J. T. (1988) Pathways and controls of the carbon cycle in salt marshes. In: Hook, D. D. et al., eds. The
ecology and management of wetlands. London, United Kingdom: Groom Helm; pp. 497-510.
Morris, J. T.; Lajtha, K. (1986) Decomposition and nutrient dynamics of litter from four species of freshwater
emergent macrophytes. Hydrobiologia 131: 215-223.
Morris, D. P.; Lewis, W. M., Jr. (1988) Phytoplankton nutrient limitation in Colorado mountain lakes.
Freshwater Biol. 20: 315-327.
Morris, J. T.; Houghton, R. A.; Botkin, D. B. (1984) Theoretical limits of belowground production by Spartina
alterniflora: an analysis through modelling. Ecol. Modell. 26: 155-175.
Morrison, I. K.; Foster, N. W. (1977) Fate of urea fertilizer added to a boreal forest Pinus bariksiana Lamb.
stand. Soil Sci. Soc. Am. J. 41: 441-448.
Mortland, M. M. (1965) Nitric oxide adsorption by clay minerals. Soil Sci. Soc. Am. Proc. 29: 514-519.
Mulder, J.; van Grinsven, J. J. M.; van Breemen, N. (1987) Impacts of acid atmospheric deposition on
woodland soils in the Netherlands: HI. aluminum chemistry. Soil Sci. Soc. Am. J. 51: 1640-1646.
Mulholland, P. J.; Palumbo, A. V.; Elwood, J. W.; Rosemond, A. D. (1987) Effects of acidification on leaf
decomposition in streams. J. N. Am. Benthol. Soc. 6: 147-158.
Munger, J. W.; Eisenreich, S. J. (1983) Continental-scale variations in precipitation chemistry. Environ. Sci.
Technol. 17: 32A-42A.
Murdoch, P. S.; Stoddard, J. L. (n.d.) The importance of nitrate in the acidification of Catskill Mountain
streams. Water Resour. Res.: Submitted.
Murray, A. J. S. (1984) Light affects the deposition of NO2 to theflacca mutant of tomato without affecting the
rate of transpiration. New Phytol. 98: 447-450.
Nadelhoffer, K. J.; Aber, J. D.; Melillo, J. M. (1985) Fine roots, net primary production, and soil nitrogen
availability: a new hypothesis. Ecology 66: 1377-1390.
August 1991
10-268 DRAFT-DO NOT QUOTE OR CITE
-------
National Atmospheric Deposition Program. (1988) NADP/NTN annual data summary. Fort Collins, CO:
National Atmospheric Deposition Program/National Trends Network Coordination Office.
National Research Council. (1969) Eutrophication: causes, consequences, correctives. Washington, DC: National
Academy of Sciences.
National Research Council. (1978) Nitrates: an environmental assessment. Washington, DC: National Academy
, of Sciences.
National Research Council. (1979) Ammonia. Baltimore, MD: University Park Press; p. 291.
National Research Council Canada. (1981) Acidification in the Canadian aquatic environment: scientific criteria
for assessing the effects of acidic deposition on aquatic ecosystems. Ottawa, ON, Canada: Environmental
Secretariat; report no. NRCC 18475.
Neely, R. K.; Davis, C. B. (1985a) Nitrogen and phosphorus fertilization of Sparganium eurycarpum Engelm.
and Typha glauca Godr. stands. I. Emergent plant production. Aquat. Bot. 22: 347-361.
Neely, R. K.; Davis, C. B. (1985b) Nitrogen and phosphorus fertilization of Sparganium eurycarpum Engelm.
and Typha glauca Godr. stands. II. Emergent plant decomposition. Aquat. Bot. 22: 363-375.
Newbold, J. D.; Elwood, J. W.; O'Neill, R. V.; Sheldon, A. L. (1983) Phosphorus dynamics in a woodland
stream ecosystem: a study of nutrient spiralling. Ecology 64: 1249-1265.
Newell, A. D.; Powers, C. F.; Christie, S. J. (1987) Analysis of data from long-terra monitoring of lakes. U. S.
Environmental Protection Agency; EPA report no. EPA-600/4-87-014.
Nicholson, K. W. (1988) The dry deposition of small particles: a review of experimental measurements. Atmos.
Environ. 22: 2653-2666.
Nihlgard, B. (1985) The ammonium hypothesis - an additional explanation to the forest dieback in Europe.
Ambio 14: 2-8.
Nilsson, J., ed. (1986) Critical loads for nitrogen and sulfur. Copenhagen, Denmark: Nordic Council of
Ministers.
Nilsson, J.; Grennfelt, P., eds. (1988) Critical loads for sulphur and nitrogen: report from a workshop; March;
Skokloster, Sweden. Copenhagen, Denmark: Nordic Council of Ministers.
Nixon, S. W. (1987) Chesapeake Bay nutrient budgets - a reassessment. Biogeochemistry 4: 77-90.
Nixon, S. W. (1988) Physical energy inputs and the comparative ecology of lake and marine ecosystems. Limnol.
Oceanogr. 33: 1005-1025.
Nixon, S. W.; Pilson, M. E. Q. (1983) Nitrogen in estuarine and coastal marine ecosystems, hi: Carpenter,
E. J.; Capone, D. G., eds. Nitrogen in the marine environment. New York, NY: Academic Press;
pp. 565-648.
Nixon, S. W.; Pilson, M. E. Q.; Oviatt, C. A.; Donaghay, P.; Sullivan, B.; Seitzinger, S.; Rudnick, D.;
Frithsen, J. (1984) Eutrophication of a coastal marine ecosystem - an experimental study using the MERL
microcosms. In: Fasham, M. J. R., ed. Flows of energy and materials in marine ecosystems: theory and
practice. New York, NY: Plenum Press; pp. 105-135. (NATO conference series. IV. Marine sciences,
v. 13).
August 1991
10-269
DRAFT-DO NOT QUOTE OR CITE
-------
Nixon, S. W.; Hunt, C. D.; Nowicki, B. L. (1986) The retention of nutrients (C, N, P), heavy metals (Mn, Cd,
Pb, Cu), and petroleum hydrocarbons in Narragansett Bay! In: Lasserre, P.; Martin, J. M., eds.
Biogeochemical processes at the land-sea boundary. Amsterdam, The Netherlands: Elsevier Science
Publishers; pp. 99-122.
Noll, K. E.; Fang, K. Y. P. (1989) Development of a dry deposition model for atmospheric coarse particles.
Atmos. Environ. 23: 585-594.
Nommik, H.; Moller, G. (1981) Nitrogen recovery in soil and needle biomass after fertilization of a Scots pine
stand, and growth response obtained. Uppsala, Sweden: Swedish University of Agricultural Sciences;
Studia Fores. Sue. Nr 159.
Norby, R. J.; Weerasuriya, Y.; Hanson, P. J. (1989) Induction of nitrate reductase activity in red spruce needles
by NO2 and HNO3 vapor. Can. J. For. Res. 19: 889-896.
Novick, N. J.; Klein, T. M.; Alexander, M. (1984) Effect of simulated acid precipitation on nitrogen
mineralization and nitrification in forest soils. Water Air Soil Pollut. 23: 317-330.
O'Dell, R. A.; Taheri, M.; Kabel, R. L. (1977) A model for uptake of pollutants by vegetation. J. Air Pollut.
Control Assoc. 27: 1104-1109.
Odum, E. P. (1971) Fundamentals of ecology. 3rd ed. Philadelphia, PA: W. B. Saunders Company; pp. 1-38,
106-136.
Odum, E. P. (1989) Ecology and our endangered life-support systems. Sunderland, MA: Sinauer Associates,
Inc.; pp. 58-66, 108-118.
Officer, C. B.; Ryther, J. H. (1980) The possible importance of silicon in marine eutrophication. Mar. Ecol.
Prog. Ser. 3: 83-91.
Officer, C. B.; Biggs, R. B.; Taft, J. L.; Cronin, L. E.; Tyler, M. A.; Boynton, W. R. (1984) Chesapeake Bay
anoxia: origin, development, and significance. Science (Washington, DC) 223: 22-27.
Ogner, O. (1972) The composition of a forest raw humus after fertilization with urea. Soil Sci. 113: 440-447.
Okano, K.; Totsuka, T. (1986) Absorption of nitrogen dioxide by sunflower plants grown at various levels of
nitrate. New Phytol. 102: 551-562.
Okano, K.; Fukuzawa, T.; Tazaki, T.; Totsuka, T. (1986) 15N dilution method for estimating the absorption of
atmospheric NO2 by plants. New Phytol. 102: 73-84. '
Okano, K.; Machida, T.; Totsuka, T. (1988) Absorption of atmospheric NO2 by several herbaceous species:
estimation by the 1SN dilution method. New Phytol. 109: 203-210.
Olsen, A. R. (1989) 1986 Wet deposition temporal and spatial patterns in North America. Richland, WA: U. S.
Department of Energy, Pacific Northwest Laboratory; report no. PNL-6933. Available from: NTIS,
Springfield, VA; DE89-014372.
Olson, R. K.; Reiners, W. A. (1983) Nitrification in subalpine balsam fir soils: tests for inhibitory factors. Soil
Biol. Biochem. 15: 413-418.
Olson, R. K.; Reiners, W. A.; Lovett, G. M. (1985) Trajectory analysis of forest canopy effects on chemical
flux in throughfall. Biogeochemistry 1: 361-373.
August 1991
10-270 DRAFT-DO NOT QUOTE OR CITE
-------
Overrent, L. ,N. (1969) Lysimeter studies on tracer nitrogen in forest soil: 2. comparative losses of nitrogen
through leaching and volatilization after the addition of urea-, ammonium- and nitrate-N15. Soil Sci.
107: 149-159. ;
Oviatt, C. A.; Keller, A. A.; Sampou, P. A.; Beatty, L. L. (1986) Patterns of productivity during
eutrophication: a mesocosm experiment.,Mar, Ecol. Prog. Ser. 28: 69-80. ,
Oviatt, C.; Lane, P.; French, F., Ill; Donaghay, P. (1989) Phytoplankton species and abundance in response to
eutrophication in coastal marine mesocosms. J. Plankton Res. 11: 1223-1244. , •
Owens, L. B.; Edwards, W. M.; Van Keuren, R. W. (1989) Sediment and nutrient losses from an unimproved,
all-year grazed watershed. J. Environ. Qual. 18: 232-238.
Pace, M. L. (1984) Zooplankton community structure, but not biomass, influences the phosphorus-chlorophyll a
relationship. Can. J. Fish. Aquat. Sci. 41: 1089-1096.
Paerl, H. W. (1985) Enhancement of marine primary production by nitrogen-enriched acid rain. Nature (London)
315: 747-749.
Paerl, H. W. (1988) Nuisance phytoplankton blooms in coastal, estuarine, and inland waters. Limnol. Oceanogr.
33: 823-847.
Paerl, H. W.; Prufert, L. E. (1987) Oxygen-poor microzones as potential sites of microbial N2 fixation in
nitrogen-depleted aerobic marine waters. Appl. Environ. Microbiol. 53: 1078-1087.
Paerl, H. W.; Crocker, K. M.; Prufert, L. E. (1987)' Limitation of N2 fixation in coastal marine waters: relative
importance of molybdenum, iron, phosphorus, and organic matter availability. Limnol. Oceanogr.
32:, 525-536. . ;
Paerl, H. W.; Rudek, J.; Mallin, M. A. (1990) Stimulation of phytoplankton production in coastal waters by
natural rainfall inputs: nutritional and trophic implications. Mar. Biol. 107: 247-254.
Parker, G. G. (1983) Throughfall and stemflow in the forest nutrient cycle. In: Macfadyen, A.; Ford, E. D.,
eds. Advances in ecological research, v. 13; pp. 57-120. ,
Parton, W. J.; Morgan, J. A.; Altenhofen, J. M.; Harper, L. A. (1988} Ammonia volatilization from spring
wheat plants. Agron. J. 80: 419-425.
Pastor, J.; Bockheim, J. G. (1984) Distribution and cycling of nutrients in an aspen-mixed-hardwood-Spodosol
ecosystem in northern Wisconsin. Ecology 65: 339-353. . • . . .
Patrick, W. H., Jr.; Delaune, R. D. (.1976) Nitrogen and phosphorus utilization by Spartina alterniflora in a salt
marsh in Barataria Bay, Louisiana. Estuarine Coastal Mar. Sci. 4: 59-64. .
Payne, W.J. (1981) Denitrification. New York, NY: Wiley Interscience. . • .. .
Peake, E.; Davidson, C. I. (1990) Wet and dry deposition of air pollutants in Alberta. In: Legge, A. H.; Krupa,
S. V., eds. Acidic deposition: sulphur and nitrogen oxides. Boca Raton, FL: Lewis Publishers, Inc.;
pp. 381-412. .- . : .
Pedrazzini, F. R.; Moore, P. A. (1983) N2O emission and changings of redox potential and pH in submerged
soil samples. Z. Pflanzenernaehr. Bodenkd. 146: 660-665.
August 1991
. 10-271
DRAFT-DO NOT QUOTE OR CITE
-------
Petenohn, W. T.; Correll, D- L. (1984) Nutrient dynamics in an agricultural watershed: observations on the role
of a riparian forest. Ecology 65: 1466-1475.
Peterjohn, W. T.; Schlesinger, W. H. (1990) Nitrogen loss from deserts in the sourthwestern United States.
Biogeochemistry 10: 67-79.
Peters, R. H. (1986) The role of prediction in limnology. Limnol. Oceatoogr. 31: 1143-1159.
Peters, N. E.; Murdoch, P. S. (1985) Hydrogeologic comparison of an acidic-lake basin with a neutral-lake basin
in the west-central Adirondack fountains, New York. Water Air Soil Pollut. 26: 387-402.
Phillips, R. A.; Stewart, K. M. (1990) Longitudinal and seasonal water chemistry variations in a northern
Appalachian stream. Water Resour. Bull. 26: 489-498.
Pietrafesa, L. J.; Janowitz, G. S.; Miller, J. M.; Noble, E. B.; Ross, S. W.; Epperly, S. P. (1986) Abiotic
factors influencing the spatial and temporal variability of juvenile fish in Pamlico Sound, North Carolina.
In: Wolfe, D. A., ed. Estuarine variability: proceedings of the eighth biennial international estuarine
research conference; July-August 1985; Durham, NH. Orlando, PL: Academic Press, Inc.; pp. 341-353.
Ponnamperuma, F. N. (1972) The chemistry of submerged soils. Adv. Agron. 24: 29-96.
Porter, L. K.; Viets, F. G., Jr.; Hutchinson, G. L. (1972) Air containing nitrOgen-15 ammonia: foliar absorption
by corn seedlings. Science (Washington, DC) 175: 759-761.
Post, W. M.; Pastor, J.; Zinke, P. J.; Stangenberger, A. G. (1985) Global patterns of soil nitrogen storage.
Nature (London) 317: 613-616.
Potts, M.; Whitton, B. A. (1977) Nitrogen fixation by blue-green algal 'communities in the intertidal zone of the
lagoon of Aldabra Atoll. Oecologia 27: 275-283.
Prairie, Y. T.; Duarte, C. M.; Kalff, J. (1989) Unifying nutrient-chlorophyll relationships in lakes. Can. J. Fish.
Aquat. Sci. 46: 1176-1182.
t * *
Prakasam, T. B. S.; Krup, M. (1982) Denitrification. J. Water Pollut. Control Fed. 54: 623-631.
Prather, R. J.; Miyamoto, S. (1974) Nitric oxide sorption by calcareous soils: UJ. effects of temperature and lack
of oxygen on capacity and rate. Soil Sci. Soc. Am. Proc. 38: 582-585.
Prather, R. J.; Miyamoto, S.; Bohn, H. L. (1973) Sorption of nitrogen dioxide by calcareous soils. Soil. Sci.
Soc. Am. Proc. 37; 860-863.
Pratt, D. M. (1965) The winter-spring diatom flowering in Narragansett Bay. Limnol. Oceanogr. 10: 173-184.
Press, M. C.; Lee, J. A. (1982) Nitrate reductase activity of Sphagnum species in the South Pennines. New
Phytol. 92: 487-494. , .
Press, M. C.; Woodin, S. J.; Lee, J. A. (1986) The potential importance of an increased atmospheric nitrogen
supply to the growth of ombrotrophic Sphagnum species. New Phytol. ,103: 45-55.
Price, K. S.; Flemer, D. A.; Taft, J. L.; Mackiernan, G. B.; Nehlsen, W.; Biggs, R. B.; Burger, N. H.;
Blaylock, D. A. (1985) Nutrient enrichment of Chesapeake Bay and its impact on the habitat of striped
bass: a speculative hypothesis. Trans. Am. Fish. Soc. 114: 97-106. .
August 1991
10-272 DRAFT-DO NOT QUOTE OR CITE
-------
Pritchett, W. L.; Comerford, N. B. (1982) Long-tenn response to phosphorus fertilization on selected
southeastern coastal plain soils. Soil Sci. Soc. Am. J. 46: 640-644.
Purchase, B. S. (1974) Evaluation of the claim that grass root exudates inhibit nitrification.. Plant .Spil'4'1:
527-539.
Quails, R. G. (1984) The role of leaf litter nitrogen immobilization in the nitrogen budget of a swamp stream.
J. Environ. Qual. 13: 640-644.
Raison, R. J.; Khanna, P. K.; Connell, M. J.; Falkiner, R. A. (1990) Effects of water availability and
fertilization on nitrogen cycling in a stand of Pinus radiata. For.' Ecol. Manage. 30: 31-43.
Rascher, C. M.; Driscoll, C. T.; Peters, N. E. (1987) Concentration and flux of solutes from snow and forest
floor during snowmelt in the most central Adirondack region of New York. Biogeochemistry.3: 209-224.
Raven, J. A. (1985) Regulation of pH and generation of osmolarity in vascular plants: a cost-benefit analysis in
relation to efficiency of use of energy, nitrogen and water. New Phytol, 101: 25-77.; .
Reddy, K. R.; Patrick, W. H. (1984) Nitrogen transformations and loss in flooded soils and sediments. Crit.
Rev. Environ. Control 13: 273-309.
Redfield, A. C. (1934) On the proportions of organic derivatives in sea Water and their relation to the
composition of plankton. In: James Johnstone memorial volume. Liverpool, United Kingdom: Liverpool
University Press; pp. 176-192.
Redfield, A. C. (1958) The biological control of chemical factors in the environment. Am. Sci. 46: 205-221.
Reed, P. B., Jr. (1988) National list of plant species that occur in wetlands: 1988 North Carolina. St. Petersburg,
FL: U. S. Fish and Wildlife Service, National Wetlands Inventory; report no. NERC-88/18:33.
Rehfuess, K. E. (1987) Perceptions on forest diseases in central Europe. Forestry 60: 1-11.
Reuss, J. O. (1983) Implications of the calcium-aluminum exchange system for -the effect of acid precipitation on
soils. J. Environ. Qual. 12: 591-595,
Reuss, J. O.; Johnson, D. W. (1986) Acid deposition and the acidification of soils and waters. New York, NY:
Springer-Verlag. (Billings, W. D.; Golley, F.; Lange, O. L.; Olson, J. S.; Remmert, H., eds. Ecological
studies: analysis and synthesis: v. 59). '
Reynolds, C. S. (1984) The ecology of freshwater phytoplankton. Cambridge, United Kingdom: Cambridge
University Press. . '
Rhee, G.-Y. (1978) Effects of N:P atomic ratios and nitrate limitation on algal growth, cell composition, and
nitrate uptake. Limnol. Oceanogr. 23: 10-25.
Rice, E. L.; Pancholy, S. K. (1972) Inhibition of nitrification by climax ecosystems. Am. J. Bot. 59: 1033-1040.
Richter, D. D.; Johnson, D. W.; Todd, D. E. (1983) Atmospheric sulfur deposition, neutralization, and ion
leaching in two deciduous forest ecosystems. J. Environ. Qual. 12: 263-270.
Riggan, P. J.; Lockwood, R. N.; Lopez, E. N. (1985) Deposition and processing of airborne nitrogen pollutants
in Mediterranean-type ecosystems of southern California. Environ. Sci. Technol. 19: 781-789.
August 1991
10-273
DRAFT-DO NOT QUOTE OR CITE
-------
Riha, S. ,T.; Campbell, G. S.; Wolfe, J. (1986) A model of competition for ammonium among heterotrophs,
nitrifiers, and roots. Soil Sci. Soc. Am. J. 50: 1463-1466. • • • .
Ro, C. U.; Tang, A. J. S.; Chan, W. H.j Kirk, R. W.; Reid, N. W.; Lusis, M. A. (1988) Wet and dry
deposition of sulfur and nitrogen compounds in Ontario. Atmos. Environ. 22: 2763-2772. ;•
Robertson, G. P.; Rosswall, T. (1986) Nitrogen in West Africa: the regional cycle. Ecol. Monogr. 56: 43-72.
Robertson, G. P.; Vitousek, P. M. (1981) Nitrification potentials in primary1 and secondary succession. Ecology
62: 376-386.
Robinson, J. B.; Whiteley, H. R.; Stammers, W.; Kaushik, N. K.; Sain, P. (1979) The fate of nitrate in small
streams and its management implications. In: Loehr, R. C.; Haith, D. A.; Walter, M. F.; Martin, C. S.,
eds. Best management practices for agriculture and silviculture: proceedings of the 1978 Cornell
agricultural waste management conference. Ann Arbor, MI: Ann Arbor Science Publishers, Inc.;
pp. 247-259.
Roelofs, J. G. M. (1983) Impact of acidification and eutrophication on macrophyte communities in soft waters in
The Netherlands: I. field observations. Aquat. Bot. 17: 139-155.
Roelofs, J. G. M. (1986) The effect of airborne sulphur and nitrogen deposition on aquatic and terrestrial
heathland vegetation. Experientia 42: 372-377.
Roelofs, J. G. M.; Schuurkes, J. A. A. R.; Smits, A. J. M. (1984) Impact of acidification and eutrophication on
macrophyte communities in soft waters. II. Experimental studies. Aquat. Bot. 18: 389-411.
Roelofs, J. G. M.; Kempers, A. J.; Houdijk, A. L. F. M.; Jansen, J. (1985) The effect of air-borne ammonium
sulphate on Pinus nigra var. maritima in the Netherlands. Plant Soil 84: 45-56.
Roelofs, J. G. M.; Boxman, A. W.; van Dijk, H. F. G. (1987) Effects of airborne ammonium on natural
vegetation and forests. In: Asman, W. A. H.; Diederen, H. S. M. A., eds. Ammonia and acidification:
proceedings [of a] symposium of the European Association for the Science of Air Pollution (EURASAP);
April; Bilthoven, The Netherlands. European Association for the Science of Air Pollution; pp. 266-276.
Rogers, H. H.; Aneja, V. P. (1980) Uptake of atmospheric ammonia by selected plant species. Environ. Exp.
Bot. 20: 251-257. ,
Rogers, H. H.; Jeffries, H. E.; Stahel, E. P.; Heck, W. W.; Ripperton, L. A.; Witherspoon, A. M. (1977)
Measuring air pollutant uptake by plants: a direct kinetic technique. J. Air Pollut. Control Assoc.
27: 1192-1197.
Rogers, H. H.; Campbell, J. C.; Volk, R. J. (1979a) Nitrogen-15 dioxide uptake and incorporation by Phaseolus
vulgaris (L.). Science (Washington, DC) 206: 333-335.
Rogers, H. H.; Jeffries, H. E.; Witherspoon, A. M. (1979b) Measuring air pollutant uptake by plants: nitrogen
dioxide. J. Environ. Qual. 8: 551-557.
Roseberg, R. J.; Christensen, N. W.; Jackson^ T. L. (1986) Chloride, soil solution osmotic potential, and soil
pH effects on nitrification. Soil. Sci. Soc. Am. J. 50: 941-945.
Rosen, K. (1982) Supply, loss and distribution of nutrients in three coniferous forest watersheds in central
Sweden. Umea, Sweden: Swedish University of Agricultural Sciences, Department of Forest Soils; report
no. 41.
August 1991
10-274 DRAFT-DO NOT QUOTE OR CITE
-------
Rosen, K. (1988) Effects of biomass accumulation and forestry on nitrogen in forest ecosystems. In: Nilsson, J.;
' Grennfelt, P., eds. Critical loads for sulphur and nitrogen: report from a workshop; March; Skokloster,
Sweden. Copenhagen, Denmark: Nordic Council of Ministers; pp. 269-293.
Rosenberg, R.; Elmgren, R.; Fleischer, S.; Jonsson, P.; Persson, G.; Dahlin, H. (1990) Marine eutrophication
case studies in Sweden. Ambio 19: 102-108.
Rosswall, T. (1981) The biogeochemical nitrogen cycle. In: Likens, G. E., ed. Some perspectives of the major
,,,,,/ .. .biogeochemical cycles. New York, NY: John Wiley & Sons; pp. 25-49. ,
Rotty, R. M. (1983) Distribution of and changes in industrial carbon dioxide production. J. Geophys. Res.
•;. ' C: Oceans Atmos. 88: 1301-1308. <
Rowland, A.; Murray, A. J. S.;.Wellborn, A. R. (1985) Oxides of nitrogen and their impact upon vegetation.
Rev. Environ. Health 5: 295-342. !
Rowland-Bamford, A. J.; Drew, M. C. (1988) The influence of plant nitrogen status on NO2 uptake, NO2
.assimilation and on the gas exchange characteristics of barley plants exposed to atmospheric NO2. J. Exp.
Bot. 39: 1287-1297.
Royal Society., (1983) The nitrogen cycle of the United Kingdom, London, United Kingdom: The Royal Society;
pp. 45-46.
Roze, F. (1988) Nitrogen cycle in Brittany heathland. Acta Oecol. Oecol. Plant. 9: 371-379.
Rudd, J. W. M.; Kelly, C. A.; Schindler, D. W.; Turner, M. A. (1988) Disruption of the nitrogen cycle in
acidified lakes. Science (Washington, DC) 240: 1515-1517.
Rudd, J. W. M.; Kelly, C. A.; Schindler, D. W.; Turner, M. A. (1990) A comparison of the acidification
efficiencies of nitric and sulfuric acids by two whole-lake addition experiments. Limnol. Oceanogr.
35: 663-679.
Rudolph, H.; Voigt, J. U. (1986) Effects of NH4+-N and NO3'-N on growth and metabolism of Sphagnum
magellanicum. Physiol. Plant. 66: 339-343.
Rydberg, L.; Edler, L.; Floderus, S.; Graneli, W. (1990) Interaction between supply of nutrients, primary
production, sedimentation and oxygen consumption in SE Kattegat. Ambio 19: 134-141.
Ryther, J. H.; Dunstan, W. M. (1971) Nitrogen, phosphorus, and eutrophication in the coastal marine
environment. Science (Washington, DC) 171: 1008-1013.
Sabey, B. R.;, Frederick, L. R.; Bartholomew,, W. V. (1959) The formation of nitrate from ammonium nitrogen
in soils: III. influence of temperature and initial population of nitrifying organisms on the maximum rate
and delay period. Soil. Sci. Soc. Am. Proc. 23: 462-465.
Sanville, W. (1988) Response of an Alaskan wetland to nutrient enrichment. Aquat. Bot. 30: 231-243.
Savant, N, K.; De Datta, S. K. (1982) Nitrogen transformations in wetland rice soils. Adv. Agron. 35: 241-302.
Saxe, H. (1986) Stomatal-dependent and stomatal-independent uptake of NOX. New Phytol. 103: 199-205.
Schaefer, D. A.; Driscoll, C. T. (n.d.) Identifying sources of snowmelt acidification with a watershed mixing
model. Water Air Soil Pollut: in press.
August 1991
10-275 DRAFT-DO NOT QUOTE OR CITE
-------
Schaefer, D. A.; Driscoll, C. T., Jr.; Van Dreason, R.; Yatsko, C. P. (1990) The episodic acidification of
Adirondack lakes during snowmelt. Water Resour. Res. 26: 1639-1647.
Schiff, S. L.; Anderson, R. F. (1987) Limnocorral studies of chemical and biological acid neutralization in two
freshwater lakes. Can. J. Fish. Aquat. Sci. 44(suppl. 1): 173-187.
Schimel, D. S.; Simkins, S.; Rosswall, T.; Mosier, A. R.; Parton, W. J. (1988) Scale and the measurement of
nitrogen-gas fluxes from terrestrial ecosystems. In: Rosswall, T.; Woodmansee, R. G.; Risser, .P.. G.,
eds. Scales and global change. New York, NY: John Wiley & Sons; pp. 179-193. "'"
Schindler, D. W. (1976) Biogeochemical evolution of phosphorus limitation in nutrient-enriched lakes of the
Precambrian Shield. In: Nriagu, J. O., ed. Environmental biogeochemistry: v. 2, metals transfer and
ecological mass balances. Ann Arbor, MI: Ann Arbor Science Publishers, Inc.; pp. 647-664.
Schindler, D. W. (1977) Evolution of phosphorus limitation in lakes. Science (Washington, DC) 195: 260-262.
Schindler, D. W. (1978) Factors regulating phytoplankton production and standing crop in the world's
freshwaters. Limnol. Oceanogr. 23: 478-486.
Schindler, D. W. (1987) Detecting ecosystem responses to anthropogenic stress. Can. J. Fish. Aquat Sci
44(suppl. 1): 6-25.
Schindler, D. W.; Paerl, H. W.; Keller.sP. E.; Lean, D. R. S. (1980) Environmental contraints onAnabaena N2
and CO2 fixation: effects of hyperoxia and phosphate depletion on blooms and chemostat cultures. In:
Hypereutrophic ecosystems: v. 2. The Hague: Junk; pp. 221-229.
Schindler, D. W.; Turner, M. A.; Hesslein, R. H. (1985) Acidification and alkalization of lakes by experimental
addition of nitrogen compounds. Biogeochemistry 1: 117-133.
Schindler, D. W.; Beaty, K. G.; Fee, E. J.; Cruikshank, D. R.; DeBruyn, E. R.; Findlay, D. L.; Linsey,
G. A.; Shearer, J. A.; Stainton, M. P.; Turner, M. A. (1990) Effects of climatic warming on lakes of
the central boreal forest. Science (Washington, DC) 250: 967-970.
Schnoor, J. L.; Palmer, W. D., Jr.; Glass, G. E. (1984) Modeling impacts of acid precipitation for northeastern
Minnesota. In: Schnoor, J. L., ed. Modeling of total acid precipitation impacts. Boston, MA:
Butterworth Publishers; pp. 155-173. (Teasley, J. L, ed. Acid precipitation series: v. 9).
Schofield, C. L.; Galloway, J. N.; Hendry, G. R. (1985) Surface water chemistry in the ILWAS basins. Water
Air Soil Pollut. 26: 403-423.
Schulze, E.-D. (1989) Air pollution and forest decline in a spruce (ficea abies) forest. Science (Washington,
DC) 244: 776-783.
Schulze, E.-D.; de Vries, W.; Hauhs, M.; Rosen, K.; Rasmussen, L.; Tamm, C.-O.; Nilsson, J. (1989) Critical
loads for nitrogen deposition on forest ecosystems. Water Air Soil Pollut. 48: 451-456.
Schuurkes, J. A. A. R. (1986) Atmospheric ammonium sulphate deposition and its role in the acidification and
nitrogen enrichment of poorly buffered aquatic systems. Experientia 42: 351-357.
Schuurkes, J. A. A. R. (1987) Acidification of surface waters by atmospheric deposition [Ph.D. thesis],
Nijmegen, The Netherlands: Catholic University.
Schuurkes, J. A. A. R.; Kok, C. J.; den Hartog, C. (1986) Ammonium and nitrate uptake by aquatic plants from
poorly buffered and acidified waters. Aquat. Bot. 24: 131-146.
August 1991
10-276 DRAFT-DO NOT QUOTE OR CITE
-------
Schuurkes, J. A. A. R.; Elbers, M. A.; Gudden, J. J. F.; Roelofs, J. G. M. (1987) Effects of simulated
ammonium sulphate and sulphuric acid rain on acidification, water quality and flora of small-scale soft
water systems. Aquat. Bot. 28: 199-226.
Schwartz, S. E. (1989) Acid deposition: unraveling a regional phenomenon. Science (Washington, DC)
243: 753-763.
Sehmel, G. A. (1980) Particle and gas dry deposition: a review. Atmos. Environ. 14: 983-1011.
Seitzinger, S. P. (1987) Nitrogen biogeochemistry in an unpolluted estuary: the importance of benthis
denitrification. Mar. Ecol.: Prog. Ser. 37: 65-73.
Seitzinger, S. P. (1988a) Denitrification in freshwater and coastal marine ecosystems: ecological and geochemical
significance. Limnol. Oceanogr. 33: 702-724.
Seitzinger, S. P. (1988b) Benthic nutrient cycling and oxygen consumption in the Delaware estuary. In:
Majumdar, S. K.; Miller, E. W.; Sage, L. E., eds. Ecology and restoration of the Delaware River Basin.
Philadelphia, PA: Pennsylvania Academy of Science; pp. 132-147.
Seitzinger, S. P.; Pilson, M. E. Q.; Nixon, S. W. (1983) Nitrous oxide production in nearshore marine
sediments. Science (Washington, DC) 222: 1244-1246.
Seitzinger, S. P.; Nixon, S. W.; Pilson, M. E. Q. (1984) Denitrification and nitrous oxide production in a
coastal marine ecosystem. Limnol. Oceanogr. 29: 73-83.
Servos, M. R.; Mackie, G. L. (1986) The effect of short-term acidification during spring snowmelt on selected
Mollusca in south-central Ontario. Can. J. Zool. 64: 1690-1695.
Setaro, F. V.; Melack, J. M. (1984) Responses of phytoplankton to experimental nutrient enrichment in an
Amazon floodplain lake. Limnol. Oceanogr. 29: 972-984.
Shaffer, P. W.; Galloway, J. N. (1982) Acid precipitation: the impact on two headwater streams in Shenandoah
'National Park, Virginia. In: Herrmann, R.; Jounson, A. L, eds. International symposium on
hydrometeorology. Denver, CO: American Water Resources Association; pp. 473-483.
Shaffer, G.; Ronner, U. (1984) Denitrification in the Baltic proper deep water. Deep Sea Res. 31: 197-220,
Sharpe, W. E.; DeWalle, D. R.; Leibfried, R. T.; Dinicola, R. S.; Kimmel, W. G.; Sherwin, L. S. (1984)
Causes of acidification of four streams on Laurel Hill in southwestern Pennsylvania. J. Environ. Qual.
13: 619-631.
Sharpe, W. E.; Leibfried, V. G.; Kimmel, W. G.; DeWalle, D. R. (1987) The relationship of water quality and
fish occurrence to soils and geology in an area of high hydrogen and sulfate ion deposition. Water
Resour. Bull. 23: 37-46.
Sharpe, W. E.; DeWalle, D. R.; Swistock, B. R. (1989) On defining acidification status of unglaciated
headwater Appalachian streams. In: Headwaters hydrology. Denver, CO: American Water Resources
Association; pp. 517-525.
Sheppard, L. J.; Smith, R. I.; Cannell, M. G. R. (1989) Frost hardiness of Picea rubens growing in spruce
decline regions of the Appalachians. Tree Physiol. 5: 25-37.
August 1991
10-277 DRAFT-DO NOT QUOTE OR CITE
-------
Shigo, A. L. (1973) Insect and disease control: forest fertilization relations. In: Leaf, A. L.; Leonard, R. E.,
eds. Forest fertilization: symposium proceedings; August 1972; Warrensburg, NY. Upper Darby, PA:
U. S. Department of Agriculture Forest Service; pp. 117-121; USDA Forest Service general technical
report NE-3. Available from: NTIS, Springfield, VA; PB-271094.
Shortle, W. C.; Smith, K. T. (1988) Aluminum-induced calcium deficiency syndrome in declining red spruce.
Science (Washington, DC) 240: 1017-1018.
Silsbee, D. G.; Larson, G. L. (1982) Water quality of streams in the Great Smoky Mountains National Park.
Hydrobiologia 89: 97-115.
Simpson, J. C.; Olsen, A. R. (1990) Wet deposition temporal and spatial patterns in North America, 1987.
Research Triangle Park, NC: U. S. Environmental Protection Agency, Atmospheric Research and
Exposure Assessment Laboratory; EPA report no. EPA-600/4-90-019. Available from NTIS, Springfield
VA; PB90-251836.
Sinclair, T. R.; van Houtte, R. F. (1982) Simulative analysis of ammonia exchange between the atmosphere and
plant communities. Agric. Environ. 7: 237-242.
Singh, H. B. (1987) Reactive nitrogen in the troposphere: chemistry and transport of NO and PAN. Environ
Sci. Technol. 21: 320-327.
Singh, J. S.; Lauenroth, W. K.; Hunt, H. W.; Swift, D. M. (1984) Bias and random errors in estimators of net
root production: a simulation approach. Ecology 65: 1760-1764.
Sinn, J. P.; Pell, E. J.; Kabel, R. L. (1984) Uptake rate of nitrogen dioxide by potato plants. J. Air Pollut.
Control Assoc. 34: 668-669.
Sisterson, D. L.; Bowersox, V. C.; Olsen, A. R.; Meyers, T. P.; Vong, R. L. (n.d.) Deposition monitoring:
methods and results. National Acid Precipitation Assessment Program; State of Science/Technology report
no. 6; in press.
Skarby, L.; Bengtson, C.; Bostrom, C.-A.; Grennfelt, P.; Troeng, E. (1981) Uptake of NOX in Scots pine. Silva
Perm. 15: 396-398.
Skeffington, R. A.; Wilson, E. J. (1988) Excess nitrogen deposition: issues for consideration. Environ Pollut
54: 159-184.
Smayda, T. J. (1974) Bioassay of the growth potential of the surface water of lower Narragansett Bay over an
annual cycle using the diatom Thalassiosira pseudonana (oceanic clone, 13-1). Limnol Oceanogr
19: 889-901.
Smith, W. H. (1974) Air pollution-effects on the structure and function of the temperate forest ecosystem
Environ. Pollut. 6: 111-129.
Smith, R. L. (1980) Ecology and field biology. 3rd ed. New York, NY: Harper & Row, Publishers; pp. 11-199.
Smith, V. H. (1982) The nitrogen and phosphorus dependence of algal biomass in lakes: an empirical and
theoretical analysis. Limnol. Oceanogr. 27: 1101-1112.
Smith, S. V. (1984) Phosphorus versus nitrogen limitation in the marine environment. Limnol. Oceanogr
29: 1149-1160.
August 1991
10-278 DRAFT-DO NOT QUOTE OR CITE
-------
Smith, C. J.; DeLaune, R. D. (1983) Gaseous nitrogen losses from Gulf Coast marshes. Northeast Gulf Sci.
6: 1-8.
Smith, C. J.; DeLaune, R. D. (1985) Recovery of added 15N-labelled ammonium-N from Louisiana Gulf Coast
estuarine sediment. Estuarine Coastal Shelf Sci. 21: 225-233.
Smith, R. E. H.; Kalff, J. (1982) Size-dependent phosphorus uptake kinetics and cell quota in phytoplankton.
J. Phycol. 18: 275-284.
Smith, V. H.; Shapiro, J. (1981) Chlorophyll-phosphorus relations in individual lakes. Their importance to lake
restoration strategies. Environ. Sci. Technol. 15: 444-451.
Smith, R. A.; Alexander, R. B.; Wolman, M. G. (1987a) Water-quality trends in the nation's.rivers. Science
' (Washington, DC) 235: 1607-1615.
Smith, R. A.; Alexander, R. B.; Wolman, M. G. (1987b) Analysis and interpretation of water quality trends in
' major U.S. rivers, 1974-1981. U. S. Geol. Surv; Water-Supply Pap. 2307.
Smullen, J. T.; Taft, J. L.; Macknis, J. (1982) Nutrient and sediment loads to the tidal Chesapeake Bay system.
In: Chesapeake Bay Program technical studies: a synthesis. Annapolis, MD: U. S. Environmental
Protection Agency; pp. 150-251. Available from: NTIS, Springfield, VA; PB84-111202.
• Sollins, P.; Grier, C. C.; McCorison, F. M.; Cromack, K., Jr.; Fogel, R.; Fredriksen, R. L. (1980) The
internal element cycles of an old-growth Douglas-fir ecosystem in western Oregon. Ecol. Monogr.
50: 261-285.
Sprent, J. I. (1987) The ecology of the nitrogen cycle. Cambridge, United Kingdom: Cambridge University
Press.
Srivastava, H. S:; Jolliffe, P. A.; Runeckles, V. C. (1975) Inhibition of gas exchanged bean leaves by NO2.
Can. J. Bot. 53: 466-474.
Steen, E. (1984) Root and shoot growth of Atriplex litoralis in relation to nitrogen supply. Oikos 42: 74-81.
Stensland, G. J.; Whelpdale, D. M.; Oehlert, G. (1986) Precipitation chemistry. In: Acid deposition: long-term
trends. Washington, DC: National Academy Press; pp. 128-199.
Steudler, P. A.; Peterson, B. J. (1984) Contribution of gaseous sulphur from salt marshes to the global sulphur
cycle. Nature (London) 311: 455-457.
Stockner, J. G.; Shortreed, K. S. (1988) Response ofAnabaena and Synechococcus to manipulation of
riitrogen:phosphorus ratios in a lake fertilization experiment. Limnol. Oceanogr. 33: 1348-1361.
Stoddard, J. L. (0000) Trends in Catskill stream water quality: evidence from historical data. Science
(Washington, DC): Submitted.
Stoddard, J. L. (1987a) Alkalinity dynamics in an unacidified alpine lake, Sierra Nevada, California. Limnol.
Oceanogr. 32: 825-839. '
Stoddard, J. L. (1987b) Micronutrient and phosphorus limitation of phytoplankton abundance in Gem Lake,
Sierra Nevada, California. Hydrobiologia 154: 103-111.
Stoddard, J. L.; Kellogg, J. H. (n.d.) Trends and patterns in lake acidification in the state of Vermont: evidence
from the Long-Term Monitoring Project. Water Air Soil Pollut.: in press.
August 1991
10-279 DRAFT-DO NOT QUOTE OR CITE
-------
Stoddard, J. L.; Murdoch, P. S. (1991) Catskill Mountains: an overview of chronic and episodic acidity in dilute
Catskill Mountain streams. In: Charles, D. F., ed. Acid deposition and aquatic ecosystems: regional case
studies of acid deposition. Springer-Verlag.
Stottlemyer, R.; Toczydlowski, D. (1990) Pattern of solute movement from snow into an upper Michigan stream.
Can. J. Fish. Aquat. Sci. 47: 290-300.
Stream Solute Workshop. (1990) Concepts and methods for assessing solute dynamics in stream ecosystems.
J. N. Am. Benthol. Soc. 9: 95-119.
Stuanes, A. O. (1980) Effects of acid precipitation on soil and forest: 5. release and loss of nutrients from a
Norwegian forest soil due to artificial rain of varying acidity. In: Drablos, D.; Tollan, A., eds.
Ecological impact of acid precipitation: proceedings of an international conference; March; Sandefjord,
Norway. Oslo, Norway: SNSF project; pp. 198-199.
Sullivan, T. J.; Christophersen, N.; Muniz, I. P.; Seip, H. M.; Sullivan, P. D. (1986) Aqueous aluminium
chemistry response to episodic increases in discharge. Nature (London) 323: 324-327.
Sullivan, T. J.; Eilers, J. M.; Church, M. R.; Blick, D. J.; Eshleman, K. N.; Landers, D. H.; DeHaan, M. S.
(1988) Atmospheric wet sulphate deposition and lakewater chemistry. Nature (London) 331: 607-609.
Sullivan, T. J.; Driscoll, C. T.; Gherini, S. A.; Munson, R. K.; Cook, R. B.; Charles, D. F.; Yatsko, C. P.
(1989) Influence of aqueous aluminium and organic acids on measurement of acid neutralizing capacity in
surface waters. Nature (London) 338: 408-410.
Summers, P. W.; Bowersox, V. C.; Stensland, G. J. (1986) The geographical distribution and temporal
variations of acidic deposition in eastern North America. Water Air Soil Pollut. 31: 523-535.
Sundaresan, B. B.; Harding, C. I.; May, F. P.; Hendrickson, E. R. (1967) Adsorption of nitrogen oxides from
waste gas. Environ. Sci. Technol. 1: 151-156.
Suttle, C. A.; Harrison, P. J. (1988) Ammonium and phosphate uptake rates, N:P supply ratios, and evidence for
N and P limitation in some oligotrophic lakes. Limnol. Oceanogr. 33: 186-202.
Swank, W. T.; Waide, J. B. (1988) Characterization of baseline precipitation and stream chemistry and nutrient
budgets for control watersheds. In: Swank, W. T.; Crossley, D. A., Jr., eds. Forest hydrology and
ecology at Coweeta. New York, NY: Springer-Verlag; pp. 57-79. (Billings, W. D.; Golley, F.; Lange,
O. L.; Olson, J. S.; Remmert, H., eds. Ecological studies: analysis and synthesis, v. 66).
Swanson, R. L.; Parker, C. A. (1988) Physical environmental factors contributing to recurring hypoxia in the
New York Bight. Trans. Am. Fish. Soc. 117: 37-47.
Swistock, B. R.; DeWalle, D. R.; Sharpe, W. E. (1989) Sources of acidic storm flow in an Appalachian
headwater stream. Water Resour. Res. 25: 2139-2147.
Switzer, G. L.; Nelson, L. E. (1972) Nutrient accumulation and cycling in loblolly pine (Pinus taeda L.)
plantation ecosystems: the first twenty years. Soil Sci. Soc. Am. Proc. 36: 143-147.
Syrett, P. J. (1953) The assimilation of ammonia by nitrogen-starved cells of Chlorella vulgaris: part I. the
correlation of assimilation with respiration. Ann. Bot. 17: 1-19.
Tallis, J. H. (1964) Studies on southern Pennine peats: III. the behaviour of Sphagnum. J. Ecol. 52: 345-353.
August 1991
10-280 DRAFT-DO NOT QUOTE OR CITE
-------
Tans, P. P.; Fung, I. Y.; Takahashi, T. (1990) Observational constraints on the global atmospheric CO2 budget.
Science (Washington, DC) 247: 1431-1438. •
Tarrant, R. F.; Miller, R. E. (1963) Accumulation of organic matter and soil nitrogen beneath a plantation of red
alder and Douglas-fir. Soil Sci. Soc. Am. Proc. 27: 231-234.
Tate, R. L., III. (1979) Effect of flooding on microbial activities in organic soils: carbon metabolism. Soil Sci.
128:267-273.
Taylor, G. E., Jr.; Hanson, P. J.; Baldocchi, D. D. (1988) Pollutant deposition to individual leaves and plant
• "canopies: sites of regulation and relationship to injury. In: Heck, W. W.; Taylor, O. C.; Tingey, D. T.,
eds. Assessment of crop loss from air pollutants. New York, NY: Elsevier Applied Science; pp. 227-257.
Taylor, B. R.; Parkinson, D.; Parsons, W. F. J. (1989) Nitrogen and lignin content as predictors of litter decay
rates: a microcosm test. Ecology 70: 97-104.
Thomas, W. H. (1970) Effect of ammonium and nitrate concentration on chlorophyll increases in natural tropical
Pacific phytoplankton populations. Limnol. Oceanogr. 15: 386-394.
Thornton, F. C.; Schaedle, M.; Raynal, D. J. (1987) Effects of aluminum on red spruce seedlings in solution
culture. Environ. Exp. Bot. 27: 489-498.
Tilman, D. (1987) Secondary succession and the pattern of plant dominance along experimental nitrogen
gradients. Ecol. Monogr. 57: 189-214.
Tiner, R. W., Jr. (1984) Wetlands of the United States: current status and recent trends. Washington, DG: U. S.
Department of the Interior, National Wetlands Inventory, Fish and Wildlife Service.
Tingey, D.-T. (1968) Foliar absorption of nitrogen dioxide [Master's thesis]. Salt Lake City, UT: University of
Utah, Department of Botany.
Tjepkema, J. D.; Cartica, R. J.; Hemond, H. F. (1981) Atmospheric concentration of ammonia in Massachusetts
and deposition on vegetation. Nature (London) 294: 445-446.
Tomlinson, G. H., II. (1983) Air pollutants and forest decline. Environ. Sci. Technol. 17: 246A-256A.
Treshow, M. (1980) Pollution effects on plant distribution. Environ. Conserv. 7: 279-286.
Triska, F. J.; Kennedy, V. C.; Avanzino, R. J.; Zellweger, G. W.; Bencala, K. E. (1990) In situ
retention-transport response to nitrate loading and storm discharge in a third-order stream. J. N. Am.
Benthol. Soc. 9: 229-239.
Tschaplinski, T. J.; Johnson, D. W.; Norby, R. J.; Todd, D. E. (n.d.) Optimum nitrogen in short rotation
sycamore plantations. Soil Sci. Soc. Am. J.: in press.
Turner, J. (1977) Effect of nitrogen availability on nitrogen cycling in a Douglas-fir stand. For. Sci.
23: 307-316.
Turner, J. (1981) Nutrient cycling in an age sequence of western Washington Douglas-fir stands. Ann. Bot.
'48: 159-169.
Turner, J.; Singer, M. J. (1976) Nutrient distribution and cycling in a sub-alpine coniferous forest ecosystem.
J. Appl. Ecol. 13: 295-301.
August 1991
10-281 DRAFT-DO NOT QUOTE OR CITE
-------
Tyler, M. (1988) Contribution of atmospheric nitrate deposition to nitrate loading in the Chesapeake Bay.
Annapolis, MD: Department of Natural Resources, Chesapeake Bay Research & Monitoring Division;
report no. AD-88-7.
Tyler, M. A.; Coats, D. W.; Anderson, D. M. (1982) Encystment in a dynamic environment: deposition of
dinoflagellate cysts by a frontal convergence. Mar. Ecol. Prog. Ser. 7: 163-178.
U. S. Environmental Protection Agency. (1982) Air quality criteria for oxides of nitrogen. Research Triangle
Park, NC: Office of Health and Environmental Assessment, Environmental Criteria and Assessment
Office; EPA report no. EPA-600/8-82-026. Available from: NTIS, Springfield, VA; PB83-131011.
U. S. Environmental Protection Agency. (1985) Ambient water quality criteria for ammonia - 1984. Washington,
DC: Criteria and Standards Division; EPA report no. EPA-440/5-85-001. Available from: NTIS,
Springfield, VA; PB85-227114.
U. S. Environmental Protection Agency. (1986) Air quality criteria for ozone and other photochemical oxidants
[draft final]. Research Triangle Park, NC: Office of Health and Environmental Assessment,
Environmental Criteria and Assessment Office; EPA report nos. EPA-600/8-84-020aF-ef. Available from-
NTIS, Springfield, VA; PB87-142949.
Ulrich, B. (1980) Production and consumption of hydrogen ions in the ecosphere. In: Hutchinson, T. C.; Havas,
M., eds. Effects of acid precipitation on terrestrial ecosystems: proceedings of the NATO conference on
effects of acid precipitation on vegetation and soils; May 1978; Toronto, ON, Canada. New York, NY:
Plenum Press; pp. 255-282.
Ulrich, B. (1983) Soil acidity and its relations to acid deposition. In: Ulrich, B.; Pankrath, J., eds. Effects of
accumulation of air pollutants in forest ecosystems: proceedings of a workshop; May 1982; Goettingen,
Federal Republic of Germany. Dordrecht, The Netherlands: D. Reidel Publishing Company;
pp. 127-146.
Urban, N. R.; Eisenreich, S. J. (1988) Nitrogen cycling in a forested Minnesota bog. Can. J. Bot. 66: 435-449.
Valiela, I.; Teal, J. M. (1974) Nutrient limitation in salt marsh vegetation. In: Ecology of halophytes. New
York, NY: Academic Press, Inc.; pp. 547-563.
Valiela, L; Teal, J. M. (1979) The nitrogen budget of a salt marsh ecosystem. Nature (London) 280: 652-656.
Valiela, L; Teal, J. M.; Sass, W. J. (1975) Production and dynamics of salt marsh vegetation and the effects of
experimental treatment with sewage sludge: biomass, production and species composition. J. Appl. Ecol
12: 973-981.
Valiela, I.; Teal, J. M.; Persson, N. Y. (1976) Production and dynamics of experimentally enriched salt marsh
vegetation: belowground biomass. Limnol. Oceanogr. 21: 245-252.
Valiela, I.; Wilson, J.; Buchsbaum, R.; Rietsma, C.; Bryant, D.; Foreman, K.; Teal, J. (1984) Importance of
chemical composition of salt marsh litter on decay rates and feeding by detritivores. Bull. Mar. Sci
35: 261-269.
van Aalst, R. M. (1982) Dry deposition of NOX. In: Schneider, T.; Grant, L., eds. Air pollution by nitrogen
oxides. Amsterdam, The Netherlands: Elsevier Scientific Publishing Company; pp. 263-270.
van Aalst, R. M.; Diederen, H. S. M. A. (1985) Removal and transformation processes in the atmosphere with
respect to SO2 and NOX. In: Zwerver, S.; van Ham, J., eds. Interregional air pollution modelling: the
state of the art. New York, NY: Plenum Press; pp. 83-103.
August 1991
10-282 DRAFT-DO NOT QUOTE OR CITE
-------
van Breemen N.; van Dijk, H. F. G. (1988) Ecosystem effects of atmospheric deposition of nitrogen in The
Netherlands. In: Dempster, J. P.; Manning, W. J.; Skeffington, R. A., eds. Excess nitrogen deposition:
[papers from the workshop]; September 1987; Leatherhead, Surrey, United Kingdom. Environ. Pollut.
54: 249-274.
van Breemen, N.; Burrough, P. A.; Velthorst, E. J.; van Dobben, H. F.; de Wit, T.; Ridder, T. B.; Reijnders,
H. F. R. (1982) Soil acidification from atmospheric ammonium sulphate in forest canopy throughtall.
. .. Nature (London) 299: 548-550. ,
van Breemen, N.; Mulder, J.; van Grinsven, J. J. M. (1987) Impacts of acid atmospheric deposition on
woodland soils in the Netherlands: H. nitrogen transformations. Soil Sci. Soc. Am. J. 51: 1634-1640.
Van der Eerden, L. J. M. (1982) Toxicity of ammonia to plants. Agric. Environ. 7: 223-235.
Van der Molen, J.; Bussink, D. W.; Vertregt, N.; van Faassen, H. G.; den Boer, D. J. (1989) Ammonia
volatilization from arable and grassland soils. In: Hansen, J. Aa.; Henriksen, K., eds. Nitrogen in
organic wastes applied to soils. New York, NY: Academic Press; pp. 185-201.
van der Valk, A. G.; Attiwill, P. M. (1983) Above- and below-ground litter decomposition in an Australian salt
marsh. Aust. J. Ecol. 8: 441-447.
van Dijk, H. F. G.; Roelofs, J. G. M. (1988) Effects of excessive ammonium deposition on the nutritional status
and condition of pine needles. Physiol. Plant. 73: 494-501.
van Dijk, H. F. G.; de Louw, M. H. J.; Roelofs, J. G. M.; Verburgh, J. J. (1990) Impact of artificial,
ammonium-enriched rainwater on soils and young coniferous trees in a greenhouse. Part II - Effects on
. - the trees. Environ. Pollut. 63: 41-59.
van Hove, L. W. A.; Koops, A. J.; Adema, E. H.; Vredenberg, W. J.; Pieters, G. A. (1987). Analysis of the
uptake of atmospheric ammonia by leaves of Phaseolus vulgaris L. Atmos. Environ. 21: 1759-1763.
van Hove, L. W. A.; Adema, E. H.; Vredenberg, W. J.; Pieters, G. A. (1989a) A study of the adsorption of
NH3 and SO2 on leaf surfaces. Atmos. Environ. 23: 1479-1486.
van Hove, L. W. A.; van Kooten, O.; Adema, E. H.; Vredenberg, W. J.; Pieters, G. A. (1989b) Physiological
effects of long-term exposure to low and moderate concentrations of atmospheric NH3 on poplar leaves.
Plant Cell Environ. 12: 899-908.
van Hove, L. W. A.; Vredenberg, W. J.; Adema, E. H. (1990) The effect of wind velocity, air temperature and
humidity on NH3 and SO2 transfer into bean leaves (Phaseolus vulgaris L.). Atmos. Environ. Part A
24: 1263-1270.
*
Van Miegroet, H.; Cole, D. W. (1984) The impact of nitrification on soil acidification and cation leaching in red
alder ecosystem. J. Environ. Qual. 13: 586-590.
Van Miegroet, H.; Cole, D. W. (1985) Acidification sources in red alder and Douglas fir soils - importance of
nitrification. Soil Sci. Soc. Am. J. 49: 1274-1279.
Van Miegroet, H.; Cole, D. W.; Homann, P. S. (1989) The effects of alder forest cover and alder forest
conversion on site fertility and productivity. In: Gessel, S. P.; Lacate, D. S.; Weetman, G. F.; Powers,
R. F., eds. Sustained productivity of forest soils: proceedings of the 7th North American forest soils
conference; July 1988; Vancouver, BC, Canada. Vancouver, BC, Canada: University of British
Columbia; pp. 333-354.
August 1991
10-283 DRAFT-DO NOT QUOTE OR CITE
-------
van Miegroet, H.; Cole, D. W.; Foster, N. W. (1991) Nitrogen distribution and cycling. In: Johnson, D. W.;
Lindberg, S. E., eds. Atmospheric deposition and nutrient cycling in forest ecosystems. New York, NY:
Springer-Verlag; in press.
Verhoeven, J. T. A.; Arts, H. H. M. (1987) Nutrient dynamics in small mesotrophic fens surrounded by
cultivated land: II. N and P accumulation in plant biomass in relation to the release of inorganic N and P
in the peat soil. Oecologia 72: 557-561.
Verhoeven, J. T. A.; Koerselman, W.; Beltman, B. (1988) The vegetation of fens in relation to their hydrology
and nutrient dynamics: a case study. In: Symoens, J. J., ed. Vegetation of inland waters. Dordrecht, The
Netherlands: Kluwer Academic Publishers; pp. 249-282. (Lieth, H., ed. Handbook of vegetation science-
v. 15/1).
Vermeer, J. G. (1986) The effect of nutrients on shoot biomass and species composition of wetland and hayfield
communities. Acta Oecol. Oecol. Plant. 7: 31-41.
Vermeer, J. G.; Berendse, F. (1983) The relationship between nutrient availability, shoot biomass and species
richness in grassland and wetland communities. Vegetatio 53: 121-126.
Verry, E. S.; Timmons, D. R. (1982) Waterborne nutrient flow through an upland-peatland watershed in
Minnesota. Ecology 63: 1456-1467.
Vince, S.; Valiela, I. (1973) The effects of ammonium and phosphate enrichments on chlorophyll a, pigment
ratio and species composition of phytoplankton of Vineyard Sound. Mar. Biol. 19: 69-73.
Vincent, W. F. (1981) Rapid physiological assays for nutrient demand by the plankton. I. Nitrogen. J. Plankton
Res. 3: 685-697.
Vitousek, P. M. (1977) The regulation of element concentrations in mountain streams in the northeastern United
States. Ecol. Monogr. 47: 65-87. -
Vitousek, P. M. (1981) Clear-cutting and the nitrogen cycle. In: Clark, F. E.; Rosswall, T., eds. Terrestrial
nitrogen cycles - processes, ecosystem strategies and management impacts: proceedings of an international
workshop; September 1979; Osterfarnebo, Sweden. Ecol. Bull. 33: 631-642.
Vitousek, P. M.; Reiners, W. A. (1975) Ecosystem succession and nutrient retention: a hypothesis. Bioscience
25: 376-381.
Vitousek, P. M.; Gosz, J. R.; Grier, C. C.; Melillo, J. M.; Reiners, W. A.; Todd, R. L. (1979) Nitrate losses
from disturbed ecosystems. Science (Washington, DC) 204: 469-474.
Vitousek, P. M.; Gosz, J. R.; Grier, C. C.; Mellilo, J. M.; Reiners, W. R. (1982) A comparative analysis of
nitrification and nitrate mobility in forest ecosystems. Ecol. Monogr. 52: 155-177.
Vollenweider, R. A. (1968) Water management research. Paris, France: OECD; report DAS/CSI/68.27.
von Liebig, J. (1840) Chemistry and its application to agriculture and physiology. London, United Kingdom:
Taylor and Walton.
Vose, J. M.; Swank, W. T. (1990) Preliminary estimates of foliar absorption of ISN labeled nitric acid vapor
mature eastern white pine (Pinus strobus L.). Can. J. For. Res.: In press.
August 1991
10-284 DRAFT-DO NOT QUOTE OR CITE
-------
Vose, J. M.; Swank, W. T.; Taylor, R. W.; Dashek, W. V.; Williams, A. L. (1989) Foliar absorption of I5N
labeled nitric acid vapor (HNO3) in mature eastern white pine (Pinus strobus L.). In: Atmospheric
deposition: proceedings of a symposium; May; Baltimore, MD. IAHS publication no. 179; pp. 4-12.
Waring, R. H. (1987) Nitrate pollution: a particular danger to boreal and subalpine coniferous forests. In:
Fujimori, T.; Kimura, M., eds. Human impacts and management of mountain forests: [proceedings of a
' symposium]. Ibaraki, Japan: Forestry and Forest Products Research Institute; pp. 93-105.
Waring, R. H.; Pitman, G. B. (1985) Modifying lodgepole pine stands to change susceptibility,to mountain pine
beetle attack. Ecology 66: 889-897.
Waring, R. H.; Schlesiriger, W. H. (1985) Forest ecosystems: concepts and management. Orlando, FL:
Academic Press, Inc.
Weetman, G. F.; Hill, S. B. (1973) General environmental and biological concerns in relation to forest
fertilization. In: Leaf, A. L.; Leonard, R. E., eds. Forest fertilization: symposium proceedings; August
1972; Warrensburg, NY. Upper Darby, PA: U. S. Department of Agriculture Forest Service; pp. 19-35;
USD A Forest Service general technical report NE-3. Available from: NTIS, Springfield, VA;
PB-271094.
Welch, E. B.; Spyridakis, D. E.; Smayda, T. (1986) Temporal chemical variability in acid sensitive high
elevation lakes. Water Air Soil Pollut. 31: 35-44.
Weller, D. E.; Peterjohn, W. T.; Goff, N. M.; Correll, D. L. (1986) Ion and acid budgets for a forested
Atlantic Coastal Plain watershed and their implications for the impacts of acid deposition. In: Correll,
D. L., ed. Watershed research perspectives. Washington, DC: Smithsonian Institution Press;
pp. 392-421.
Wells, C. G. (1971) Effects of prescribed burning on soil chemical properties and nutrient availability. In:
; Prescribed burning symposium: proceedings; April; Charleston, SC. Asheville, NC: U. S. Department of
Agriculture Forest Service, Southeastern Forest Experiment Station; pp. 86-99.
Wesely, M. L.; Eastman, J. A.; Stedman, D. H.; Yalvac, E. D. (1982) An eddy-correlation measurement of
NO2 flux to vegetation and comparison to O3 flux. Atmos. Environ. 16: 815-820.
West, N. E.; Skujins, J., eds. (1978) Nitrogen in desert ecosystems. Stroudsburg, PA: Dowden, Hutchinson &
Ross, Inc. (US/IBP synthetic series 9).
Westerman, R. L.; Tucker, T. C. (1978) Denitrification in desert soils. In: West, N. E.; Skujins, J., eds.
' Nitrogen in desert ecosystems. Stroudsburg, PA: Dowden, Hutchinson & Ross, Inc.; pp. 75-106.
(US/IBP synthesis series 9). '
Westman, W. E. (1977) How much are nature's services worth? Science (Washington, DC) 197: 960-964.
Wetselaar, R.; Farquhar, G. D. (1980) Nitrogen losses from tops of plants. Adv. Agron. 33: 263-302.
Wetzel, R. G. (1983) Limnology. 2nd ed. Philadelphia, PA: W. B. Sanders.
Wheeler, B. D.; Giller, K. E. (1982) Species richness of herbaceous fen vegetation in Broadland, Norfolk in
relation to the quantity of above-ground plant material. J. Ecol. 70: 179-200.
White, T. C. R. (1984) The abundance of invertebrate herbivores in relation to the availability of nitrogen in
stressed food plants. Oecologia 63: 90-105.
August 1991
10-285
DRAFT-DO NOT QUOTE OR CITE
-------
Whitford, L. A.; Schumacher, G. J. (1961) Effect of current on mineral uptake and respiration by a fresh-water
alga. Limnol. Oceanogr. 6: 423-425. . •
Whitford, L. A.; Schumacher, G. J. (1964) Effect of a current on respiration and mineral uptake in Spirogyra
and Oedogonium. Ecology 45: 168-170.
Wiebe, W. J.; Johannes, R. E.; Webb, K. L. (1975) Nitrogen fixation in a coral reef community. Science
(Washington, DC) 188: 257-259.
Wigington, P. J.; Davies, T. D.; Tranter, M.; Eshleman, K. (1989) Episodic acidification of surface waters due
to acidic deposition. Washington, DC: National Acid Precipitation Assessment Program; State-of-science
technology report 12.
Williams, M. W.; Melack, J. M. (n.d.a) Precipitation chemistry and ionic loading to an alpine basin, Sierra
Nevada. Water Resour. Res.: In press.
Williams, M. W.; Melack, J. M. (n.d.b) Solute chemistry of snowmelt and runoff in an alpine basin, Sierra,
Nevada. Water Resour. Res.: in press.
Winner, W. E.; Atkinson, C. J. (1986) Absorption of air pollution by plants, and consequences for growth.
TREE 1: 15-18.
Wisheu, I. C.; Keddy, P. A. (1989) The conservation and management of a threatened coastal plain plant
community in eastern North America (Nova Scotia, Canada). Biol. Conserv. 48: 229-238.
Woodin, S. J.; Lee, J. A. (1987) The fate of some components of acidic deposition in ombrotrophic mires.
Environ. Pollut. 45: 61-72.
Woodin, S.; Press, M. C.; Lee, J. A. (1985) Nitrate reductase activity in Sphagnum fuscum in relation to wet
deposition of nitrate from the atmosphere. New Phytol. 99: 381-388.
Woodmansee, R. G. (1978) Additions and losses of nitrogen in grassland ecosystems. Bioscience 28: 448-453.
Woodwell, G. M. (1970) Effects of pollution on the structure and physiology of ecosystems: changes in natural
ecosystems caused by many different types of disturbances are similar and predictable. Science
(Washington, DC) 168: 429-433.
World Health Organization. (1987) The effects of nitrogen on vegetation. In: Air quality guidelines for Europe.
Copenhagen, Denmark: Regional Office for Europe; pp. 373-385. (WHO regional publications, European
series no. 23).
Worsnop, G.; Will, G. M. (1980) Fate of 15N urea fertiliser applied to a recently thinned radiata pine stand on a
pumice soil. N. Z. J. For. Sci. 10: 381-394.
Wulff, F.; Stigebrandt, A.; Rahm, L. (1990) Nutrient dynamics of the Baltic Sea. Ambio 19: 126-133.
Wurtsbaugh, W. A.; Home, A. J. (1983) Iron in eutrophic Clear Lake, California: its importance for algal
nitrogen fixation and growth. Can. J. Fish. Aquat. Sci. 40: 1419-1429.
Yates, P.; Sheridan, J. M. (1983) Estimating the effectiveness of vegetated floodplains/wetlands as nitrate-nitrite
and orthophosphorus filters. Agric. Ecosyst. Environ. 9: 303-314.
August 1991
10-286
DRAFT-DO NOT QUOTE OR CITE
-------
Yoch, D. C.; Whiting, G. J. (1986) Evidence for NH4+ switch-off regulation of nitrogenase activity by bacteria
in salt marsh sediments and roots of the grass Spartina alterniflora. Appl. Environ. Microbiol.
51: 143-149.
Zemba, S. G.; Golomb, D.; Fay, J. A. (1988) Wet sulfate and nitrate deposition patterns in eastern North
America. Atmos. Environ. 22: 2751-2761.
•August 1991
10-287
DRAFT-DO NOT QUOTE OR CITE
-------
-------
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
11. EFFECTS OF NITROGEN OXIDES
ON VISIBILITY *
Clear days are an important aesthetic resource for us all. They also carry
commercial value for tourism and real estate. Thus, the appearance of layers of
smoggy haze over cities and across rural vistas is one of the most widely noticed
effects of air pollution (Sloane and White, 1986).
Emissions of nitrogen Oxides (NOX) can contribute significantly to visibility impairment,
or the "layers of smoggy haze" noted by Sloane and White. They Can have aesthetic impact
because they can cause a yellow-brown discoloration of the atmosphere when present in
plumes or in urban, regional, and layered haze. They can also reduce visual range, thereby
diminishing the contrast of distant objects viewed through an atmosphere containing nitrogen
oxides.
Only some of the species in the NOX family, however, are optically active and thus able
to affect atmospheric visibility. Figure 11-1 illustrates the major categories (including
atmospheric oxidation products) of NOX species and the two species that have an effect on
visibility: nitrogen dioxide (NO^, a gas that absorbs light, chiefly at the blue end of the
visible spectrum; and nitrate aerosols (NOg), particles that scatter light. The other forms of
NOX that occur in ambient air, nitric oxide (NO), nitrous acid (HNO2), and nitric acid
(HNO3), are optically inactive gases and therefore do not contribute to visibility impairment.
(Nitrous acid, however, interferes with chemiluminescence NO2 measurements and therefore
would indirectly affect the estimation of the effects of NO2 on visibility.). Thus, depending
on the form in which NOX exists in the atmosphere, NOX may or may not play a significant
role overall in visibility. For example, nitrate aerosol may never form from nitric acid in
warm climates; in areas with low ambient atmospheric concentrations of ammonia (NH3); or
in areas with high ambient concentrations of sulfate (SOJ), since sulfate preferentially reacts
with available atmospheric ammonia.
Nitrogen oxides have been found to play a significant role in the aesthetic impact caused
by combustion emission sources such as power plants. This impact is dominated by the
yellow-brown coloration caused by NO2 relatively near the source (within 100 km). Nitrate
aerosols have been found to play a significant role in the haze observed in urban areas;
August 1991
11-1
DRAFT-DO NOT QUOTE OR CITE
-------
&<+>+v
a*i:|:|:;:;:;
I
•ST.
.
ili
LLI
O
g
o
•
(0
o-
il*
o~
I
4)
I
T3
8
1
I
•s
(U
I
TH
rH
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
particularly during winter and near significant ammonia sources (such as cattle grazing fields
and feedlots). Nitrate aerosols, along with sulfate, may also play a significant role in the
formation of wintertime layered haze that has been observed in the vicinity of large, isolated
power plants.
Although NOX has a clearly defined effect on visibility (aesthetic impacts and visual
range reduction), in most areas of the country visibility impairment is usually dominated by
other species, such as sulfate and elemental and organic carbon particles. Also, it should be
noted that brownish atmospheric discoloration may be caused by non-nitrogenous particles
such as sulfate and not solely by NO2 and nitrate.
11.1 OVERVIEW OF LIGHT SCATTERING AND ABSORPTION
The visibility effects of the optically active forms of NOX, nitrogen dioxide (NO2) and
nitrate aerosols (NO^), can best be illustrated by reviewing some of the fundamentals of
atmospheric optics. The deterioration of visibility is the result of the absorption and
scattering of light by gaseous molecules and suspended solid or liquid particles (Middleton,
1952). Absorbed light is transformed into other forms of energy, such as heat, while
scattered light is re-radiated in all directions.
The effect of the intervening atmosphere on the visibility and coloration of a viewed
object, such as the horizon sky, a distant mountain-, or a cloud, can be calculated by solving
the radiation transfer equation along the line of sight (see schematic in Figure 11-2). This
equation can be solved if the light extinction properties of the intervening atmosphere are
known.
The change in the light intensity of a specific wavelength, or spectral radiance I(A), as a
function of distance along the line of sight can be calculated as follows (Chandrasekhar,
1960; Larimer and Samuelsen, 1975, 1978; Latimer et al., 1978; White et al., 1986):
;= -bext(A)I(A)
4?r
(11-1)
28
August 1991
11-3
DRAFT-DO NOT QUOTE OR CITE
-------
SUN
ELEMENTAL VOLUME
<
1 Figure ]
2
3 Source: I
4
5
6
7 where
8
9 I(A)
10
11 r
12
13
14 p(A,9)
15
16 FS(A)
17
18 D
19
20
21
22 bscat(A)
23
24 bext(A)
August 1
\ (CONTAINING AIR,
\ / PARTICLES, AND NO2)
SCATTERING /*\ /
ANGLE, 6 / \ t. LINE OF SIGHT
/ / x
*..„..> _ fc . . / iV
y ^r^ *• — ••••.• l>
/ V j , jj
i — t a
OBJECT OBSERVER
11-2. Schematic of an elemental volume of haze along a line of sight.
.atimer and Ireson (1980).
= the spectral light intensity of wavelength A,
= the distance along the line of sight from the object to the observer (see
Figure 11-2 for definitions),
= the scattering distribution or phase function for the scattering angle 6,
= the solar flux (watt/m2/jiim) incident on the line of sight,
= the diffuse light source term. (Diffuse light emanates from directions other tl
that of the sun. This light results from light scattered from other portions of
atmosphere and light reflected from the surface of the Earth and from clouds.
= the light scattering coefficient, and
= the light extinction coefficient, the sum of scattering and absorption.
991 n_4 DRAFT-DO NOT QUOTE OR C
.•>
lan
the
)
ITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
An examination of Equation 11-1 indicates that light can be both removed and added to
the line of sight. The first term on the right side of this equation represents the rate at which
light is removed from the line of sight, while the second term is the rate at which it is added.
If the first term is larger than the second, the net effect is a decrease in light intensity
(darkening) of an observed object as one moves along the line of sight (see upper curve in
Figure 11-3). If the second term is larger than the first, the net effect is an increase in light
intensity (brightening) of an observed object. The darkening effect, the first term, is
dependent on total light extinction (bext), which is the sum of light scattering and absorption.
The brightening effect, the second term, is dependent only on light scattering (bscat). Thus,
light absorption can only darken objects viewed through the atmosphere, while light scattering
can either brighten or darken viewed objects. Since NO2 is a gas that preferentially absorbs
blue light, it always tends to darken and discolor the sky and objects viewed through the
atmosphere. Since nitrate aerosol scatters light, it can either brighten or darken the sky and
objects. In Equation 11-1, the light extinction coefficient is the sum of its light scattering and
light extinction components:
= bscat(X) + babs(X) = (bsg + bsp) + (bag + bap).
(11-2)
The first term, b , is the scattering coefficient attributable to gases and is the result
primarily of Rayleigh scattering caused by gases in the atmosphere (chiefly N2 and O^). The
second term, b , is the scattering coefficient from particles suspended in the atmosphere
(aerosols). Nitrate aerosol contributes to this term, along with other aerosols, including
sulfates, organic and elemental carbon, and other particulate matter, both fine (< 2.5 ^m in
diameter) and coarse (> 2.5 ^m in diameter). The third term, bag, is the absorption
coefficient resulting from gases. Nitrogen dioxide is the only significant contributor to this
term in the visible spectrum. The fourth and last term, bap, is the absorption coefficient
resulting from particles. This term is dominated by the effect of elemental carbon (soot), a
combustion product found, for example, in diesel engine exhaust.
All of these components of total light extinction, as well as total extinction itself, are
functions of the wavelength of light. As discussed in more detail later, the atmospheric
August 1991
11-5
DRAFT-DO NOT QUOTE OR CITE
-------
6-1
BRIGHT OBJECT
UQHT INTENSITY OF HORIZON
BLACK OBJECT
OBJECT-OBSERVER DISTANCE, re
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
Figure 11-3. Effect of a homogeneous atmosphere on light intensity of bright and dark
objects as a function of distance along a line of sight.
Source: Latimer and Ireson (1980).
discoloration caused by NOX (both NO2 and nitrate aerosol) can be explained by the
wavelength-dependent nature of NO2 light absorption and nitrate light scattering effects.
Both scattering and absorption from these NOX species are stronger at the blue end of the
visible spectrum (wavelength A = 0.4 /tin) than at the red end (X = 0.7 /*m).
The light extinction (bext) coefficient is the optical equivalent of ambient pollutant
concentration. This parameter (as well as its scattering and absorption components) has units
of inverse distance (e.g., m"1, km"1, Mm"1). These coefficients can be considered to be the
equivalent light extinction, scattering, or absorption cross-sectional area (m2) per unit volume
of ambient air (m3). The scattering or absorption coefficient can be determined from the
product of the concentration of an optically active species and its light scattering or specific
August 1991
11-6
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
absorption efficiency (/8). This efficiency is commonly stated in units of m2/g. When the
ambient concentration Gig/m3) of a given species is multiplied by its extinction efficiency
(m2/g), the extinction coefficient of that species, in units of inverse megameters (Mm"1), is
obtained.
The light extinction efficiency for particles (absorption, bap, and especially scattering,
bsp) is a strong function of particle size (see Figure 11-4). Fine particles, those with
diameters less than 2.5 /mi, are much more effective per unit mass in scattering light than are
coarse particles, those with diameters > 2.5 jum. Particle scattering efficiency is a maximum
0.
TYPICAL
NONABSORB1NG
AEROSOL
10.0
PARTICLE DIAMETER,
11
12
13
14
Figure 11-4. Light extinction efficiency at A = 0.55 /tin as a function of particle size for
soot and for typical, nonabsorbing atmospheric aerosol.
Source: Latimer (1988a) after Bergstrom (1973). "
August 1991
11-7
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
for particles having a diameter of approximately 0.5 jwm. Coarse particles have scattering
efficiencies that are approximately an order of magnitude smaller (see Figure 11-4).
Nitrate particles can be either coarse or fine. Milford and Davidson (1987) reviewed
the sizes of particulate sulfate and nitrate in the atmosphere; nitrate mass median diameters
ranged from 0.23 to 4.2 /*m in 16 different measurement sets. Wolff (1984) noted that in
continental environments nitrate can exist as either coarse or fine; however, in a number of
summertime studies in the eastern U.S., nitrate concentrations were quite low and nitrate
occurred primarily in the coarse mode (Wolff, 1984; Mamane and Dzubay, 1986). Wolff
explained this qualitatively by the reaction of alkaline soil dust with HNO3; nitrate aerosol is
not formed in the submicron mode if temperatures are high or ammonia is not available or is
tied up with sulfate. It should be noted, however, that the data of Wolff (1984) were
collected using methods later found to have significant artifact problems. In coastal
environments, nitrate may also be primarily in the coarse mode because of reaction with sea
salt (Yoshizumi, 1986; Wall et al., 1988; Orel and Seinfeld, 1977; Mamane and Mehler,
1987). Richards (1983) suggested that coarse particle nitrate may form from nighttime
oxidation involving N2O5-H2O reactions on the surfaces of particles. Nitrate is in the
submicron fine mode when it reacts directly with ammonia to form ammonium nitrate
(Orel and Seinfeld, 1977; Wolff, 1984). The submicron nitrate forms when conditions are
favorable (abundant ambient ammonia and moderate temperatures).
Nitrate aerosol in the size range of 0.1 to 2.5 pm is most effective per unit mass in
scattering light. For particles having a typical density (/>) of 2 g/cm3 and a diameter of
0.5 jtm, Figure 11-4 shows that the scattering efficiency at the middle of the visible spectrum
(X = 0.55 urn) is approximately 5 m2/g. By contrast, the average NO2 absorption efficiency
over the wavelengths 0.45 to 0.65 /^m, centered on 0.55 jt*m, is 0.144 m2/g (Latimer and
Ireson, 1988, based on Dixon, 1940). Thus, the extinction efficiency of nitrate aerosol can
be more than an order of magnitude greater than that for NO2. As discussed in the next
section, the extinction efficiencies of both nitrate aerosol and NO2 gas are strong functions of
the wavelength, being larger at the blue end (X = 0.4 /*m) of the visible spectrum.
August 1991
11-8
DRAFT-DO NOT QUOTE OR CITE
-------
1 11.2 ATMOSPHERIC DISCOLORATION CAUSED BY NITROGEN
2 OXIDES
3 As Finlan (1981) so aptly stated: "Many of the most beautiful sights in nature are
4 caused by wavelength-dependent light scattering. It can be truly exhilarating to see the
5 beauty of the blue sky or to witness a rainbow, a sunset or a sunrise. Unfortunately, the
6 physical processes responsible for these beautiful sights also cause much of the color that we
7 often see in smogs and hazes over cities."
8 The undesirable yellow or whisky-brown color of hazes has been an ongoing topic of
9 discussion in the literature for more than 20 years. Hodkinson (1966) described the effects
10 that NO2 could produce on the color of the atmosphere. Charlson and Ahlquist (1969),
11 however, argued that wavelength-dependent scattering was the primary cause of atmospheric
12 discoloration in most situations. Horvath (1971) countered with the argument that any color
13 caused by wavelength-dependent light scattering that removed light from the line of sight
14 would be offset by the additional light scattered into the line of sight by the same wavelength-
15 dependent scattering. Thus, he thought that any color would be the result of the absorption
16 of blue light by NO2. He did conclude, however, that if extremely bright objects were
17 viewed through an aerosol, a discoloration could result. Charlson et al. (1972) measured
18 NO2 concentrations and the wavelength dependence of the light scattering coefficient in
19 Pasadena, California, during August and September 1970 and concluded that NO2 had a
20 significant effect on atmospheric color 20 percent of the time. Sloane (1987) applied Mie
21 theory to calculate the effects of urban haze mixtures of NO2 and elemental carbon (soot).
22 She found that soot can offset the coloration caused by NO2, even though both species absorb
23 preferentially at the blue end of the spectrum. Husar and White (1976) performed careful
24 atmospheric optics calculations using Mie scattering theory (Kerker, 1969) to assess the
25 relative roles of wavelength-dependent light scattering by particles and wavelength-dependent
26 light absorption caused by NO2. They found that particles typical of Los Angeles haze could
27 cause yellow-brown discoloration when the sun was behind the observer (scattering angle
28 8 > 90°), and typical NO2 concentrations could perceptibly add to this color. More detailed
29 analysis by Finlan (1981) confirmed the importance of scattering angle and the size
30 distribution and refractive index of the aerosol in determining atmospheric color.
August 1991
11-9
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
Atmospheric color can be studied theoretically by solving Equation 11-1 for the spectral
radiance or light intensity of an object observed at distance r as follows (Middleton, 1952;
Latimer and Samuelsen, 1975, 1978; Latimer et al., 1978; Husar and White, 1976; White
etal., 1986):
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
Ir(A) = I0(A) exp(-r) + J [1 - exp(-r)],
(11-3)
where
Ir, I0 = spectral light intensities at distance r from an object and at the object itself,
r = optical depth between the object and the observer (= J bext dr),
J = the source function (the second term in Equation 11-1, divided by bext).
Equation 11-3 can be used to evaluate the effect of a uniform concentration of NO2 on
atmospheric coloration. The ratio of the intensity of the horizon sky (h) with and without a
given concentration of NO2 can be calculated from Equation 11-3 as follows (Hodkinson,
1966; Robinson, 1968; White, 1982):
= (1 + b V
(11-4)
23
24
25
26
27
28
29
30
31
32
33
The light absorption coefficient for NO2, bag, is a stiong function of wavelength.
Figure 11-5 shows the wavelength dependence of the NO2 light absorption coefficient over
the visible spectrum (Dixon, 1940; Latimer and Ireson, 1980); The value at the blue end of
the visible spectrum, X = 0.4 jitm, is 1.71 km"1 ppm"1, more than five .times larger than the
value at the center of the visible spectrum at a green wavelength A = 0.55 jttm (0.33 km"1
ppm"1), and approximately 100 times larger than at the red end of the spectrum at a
wavelength of 0.7 /*m (0.017 km"1 ppm"1). When Equation 11-4 is evaluated as a function
of wavelength (A), the curves shown in Figure 11-6 are obtained for the horizon sky light
intensity ratio (Hodkinson, 1966; White, 1982). Nitrogen dioxide causes a darkening effect,
especially at the blue end of the visible spectrum. For example, with an NO2-visual range
August 1991 H_10 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
§
E
> 0.05
J 0.04
0.01
0.4
0.5 0.6
WAVELENGTH (A), nm
Figure 11-5. Light absorption efficiency of nitrogen dioxide as a function of the
wavelength of light hi the visible spectrum, 0.4 jon < A < 0.7 jon.
Source: Latimer and Ireson (1980) after Dixon (1940).
product of 0.3 ppm-km, the horizon sky light intensity at A = 0.4 /Ltm is about 14 percent
less than it would be without NO2 and would thus be quite noticeably discolored (yellow or
^
brown). This concentration-visual range product could be caused by 0.03 pp'm (60 /*g/m )
NO2 associated with a visual range of 10 km (typical of urban haze) and by 0.0015 ppm
(3 jug/m3) NO2 associated with a visual range of 200 km (typical of much of the nonurban
western U.S.).
Atmospheric aerosols, including particulate nitrates, can also cause atmospheric
discoloration (Ahlquist and Charlson, 1969; Husar and White, 1976). The scattering
Month 1991
11-11
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
0.4
BLUE
0.5 0.6
WAVELENGTH, urn
O.7
RED
Figure 11-6. Effect of nitrogen dioxide on horizon sky brightness as a function of the
wavelength of light; relative horizon brightness, bext/(bext + b ) for
selected values of the product of NO2 concentration and visual range
assuming that b&it = 3/(visual range).
Source: White (1982) adapted from Hodkinson (1966).
coefficient of particles smaller .than 1.5 /tm in diameter can be strongly dependent on the
wavelength of light as shown in Table 11-1 (Latimer and Ireson, 1980). For example, an
aerosol with a mass median diameter of 0.5 ^m has a light scattering coefficient b f which
SCcil
is inversely proportional to wavelength X. Thus, light scattering at the blue end
(A = 0.4 ^m) of the visible spectrum would be 75 percent greater (7/4 = 1,75) than at the
red end (A = 0.7 jim). Since the light scattering coefficient caused by aerosols and the light
absorption coefficient caused by NO2 are both wavelength dependent, both can cause
atmospheric discoloration.
August 1991
11-12 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
TABLE 11-1. WAVELENGTH DEPENDENCE OF LIGHT SCATTERING
COEFFICIENT AS A FUNCTION OF PARTICLE SIZE DISTRIBUTION
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
Mass Median
Diameter
0
0
0
0
0
0
(DG),a/im
.1
.2
.3
.4
.5
.6
0.8
1
.0
>5
cP
2.8
2.1
1.6
,.-1.2
1.0
0.7
0.5
0.2
0
"Geometric standard deviation crg = 2.
ba is defined as follows:
[A
b nfLA-i) = 0 JXj) —
bL-dU *• JL' ovdV- £j \
A
i
± -a
2
(appropriate for 0.4 < A < 0.7
Source: Latimer and Ireson (1980).
Husar arid White (1976)" formulated the problem of atmospheric coloration rigorously in
terms of radiation transfer theory. A solution was derived from theory and from aerosol size
distributions measured in Los Angeles. They found that aerosol (without NO^ could cause
yellow-brown discoloration, and that this discoloration would increase as NO2 concentrations
increase and as the scattering angle 9 increases. Noticeable discoloration from NO2 was
found to occur at concentrations as low as 0.05 ppm. The discoloration effect caused by
particles, unlike that caused by NO2, is dependent on the scattering angle, B, with most
intense effects occurring in situations in which the sun is behind the observer (8 > 90°). In
addition, when the,viewed object has a light intensity greater than the horizon sky light
intensity (the Ih asymptote in Figure 11-3), light, scattered by fine particles would cause a
darkening and discoloring effect because of the wavelength-dependent light scatter.
Waggoner et al. (1983) used teleradiometer measurements to determine the color of the
winter haze in Denver that is commonly known as the "brown cloud." Although this haze
appeared to be brown in contrast to the blue sky above, they found that its spectral .
light-intensity distribution was gray and was caused primarily by aerosol .rather than NO2.
August 1991
11-13
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
These findings were consistent with the conclusions of Horvath (1971) and Husar and White
(1976) that yellow haze could appear brown if it were darker than the viewing background.
The chromatic adaptation of the human eye-brain system (Cornsweet, 1970) also explains
why a gray haze may appear yellow or brown. An observer that has adapted to the color of
the blue sky will visually perceive a gray haze as the complementary color to that adaptation
(i.e., yellow or brown).
11.3 VISUAL RANGE REDUCTION CAUSED BY NITROGEN OXIDES
At some distance from a black object, an observer can no longer distinguish between the
intensity of it and the sky. This limit of perceptibility is defined by a threshold (liminal)
contrast which is just noticeable to a human observer. The distance at which the contrast of a
black object against the horizon sky equals this threshold is called the visual range, or
commonly, visibility. Although a range of values for the threshold contrast from about 1 to
5 percent is supported by the literature (Middleton, 1952; U. S. Environmental Protection
Agency, 1979; Latimer, 1988b), the threshold human visual perception threshold is
commonly assumed to be a contrast of 2 percent. «
Koschmieder (1924) developed a formula for visual range, which is based on the
assumptions that the threshold contrast is 2 percent, that the atmosphere is uniform and cloud-
free, and that the curvature of the Earth can be ignored when evaluating horizon light
intensity. The Koschmieder equation is simply:
rv = - ln(Cmin)/b(
(H-5)
where
ext
the visual range,
the contrast perceptibility threshold, and
the light extinction coefficient, as
defined previously.
If the commonly accepted threshold of 2 percent is used above, the Koschmieder equation
becomes
rv = 3.9/bext,
(11-6)
August 1991
11-14 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
the most common form of. the equation. If the perceptibility threshold is assumed to be
5 percent, which appears to correlate best with common airport visibility measurements
(Johnson, 1981; Latimer, 1988b), the equation becomes:
rv = 3/bexf
(11-7)
Note that as the light extinction coefficient increases, visual range decreases. This
inverse relationship suggests that increases in atmospheric concentrations of light scattering
and absorbing species will cause a decrease in visibility. Figure 11-7 illustrates this
/^
relationship for fine particles assumed to have a scattering efficiency of 4 m /g (U.S.
Environmental Protection Agency, 1979). Since both of the optically active NOX species,
NO2 and nitrate aerosol, contribute to the absorption and scattering components of light
extinction (bext), they both tend to reduce visual range. -,.,....
If it is not uniformly distributed in the atmosphere, NO2 may not contribute to a
reduction in the contrast of a distant object and hence to visual range reduction. This can
happen when NO2 is located relatively close to the observer (e.g.,,in a plume or haze layer).
In such a situation, the light absorbed by NO2 reduces the light intensity of both the sky and.
the dark object equally, so that the sky and object are darkened but their contrast remains
unaffected. Latimer and Samuelsen (1975, 1978) developed a formula to account for this
effect for atmospheres containing NO2 plumes.
11.4 NITRATE PHASE CHANGES AND HYGROSCOPICITY
The role played by nitrate particles in urban, regional, and layered haze and in plumes
is currently uncertain because of the volatile nature of this species. Unlike sulfate, which is
always in the particulate phase, nitrate often remains in the gas phase as nitric acid. In order
for condensation of particulate nitrate (ammonium nitrate, NH4NO3) to occur, there must be
sufficient atmospheric ammonia to react with nitric acid. Furthermore, the vapor pressure of
ammonium nitrate is strongly temperature-dependent, so that even if ammonia is present in
the atmosphere nitrate particles may not condense because of moderate or high temperatures.
August 1991
11-15
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
'9
10
11
12
13
14
15
16
^ADDITION OF 1/
-------
1 The issue of changes in phase between gas and aerosol is a key uncertainty in
2 understanding, measuring, and mathematically modeling the impacts of nitrate aerosol (Sloane
3 and White, 1986):
4 Just as a cloud produces a dramatic visual effect when only a small fraction of the water
5 vapor changes phase, a substantial naze results if only a fraction of the gaseous pollutant
6 mass enters a condensed phase. In this regard, visibility is unique among air pollution
7 effects; it depends not only on the amount of air pollution but in addition on its phase.
8 This peculiarity greatly complicates the prediction of visibility impairment and aerosol
9, measurement procedures because the equilibrium between the condensed and gaseous
10 phases can be fragile.
12 Ammonium nitrate particles will form only if (1) sufficient ambient ammonia is present
13 to neutralize gas phase nitric acid (HNO3) and (2) temperatures and relative humidities are
14 such that the thermodynamic equilibrium favors the formation of nitrate aerosol (Stelson
15 et al., 1979; Stelson and Seinfeld, 1982; Saxena et al., 1986; Sloane and White, 1986).
16 Ammonia reacts preferentially with acidic sulfate compounds until it is fully neutralized as
17 ammonium sulfate (Saxena et al., 1986). If sufficient gas phase ammonia is left after sulfate
18 neutralization and temperatures are low enough, ammonium nitrate aerosol will condense. At
19 relative humidities above 62 percent, the deliquescent point for ammonium nitrate, water
20 vapor is taken up in the nitrate particle (droplet), forming a water solution (Saxena et al.,
21 1986). At these higher relative humidities, a new equilibrium is established favoring more
22 nitrate in the particulate phase (Sloane and White, 1986).
23 The net result of all of the nitrate phase interactions is that particulate ammonium nitrate
24 "can build up only in locations where sufficient ammonia is present to neutralize the sulfuric
25 acid. This occurs, for example, in Los Angeles and Denver, where sulfate concentrations are
26 relatively low compared to concentrations of ammonia" (Milford and Davidson, 1987).
27 White and Macias (1987) attribute the extremely low nitrate aerosol concentrations observed
28 in the intermountain West to very low ambient nitric acid and ammonia concentrations and to
29 the warm temperatures during the non-winter months. Thus, the conditions can be
30 summarized under which fine nitrate particles are most likely to form: high ambient
31 concentrations of ammonia and nitric acid (e.g., Los Angeles, Denver), low ambient
32 concentrations of sulfate (e.g., mo'st of the western U.S.), low temperatures (e.g., winter),
33 and high humidities (e.g., winter, coastal sites). Conversely, fine nitrate particles are least
34 likely to form under the following conditions: low ambient concentrations of ammonia and
August 1991
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
nitric acid (e.g., intermountain West), high ambient concentrations of sulfate
(e.g., the eastern U.S.), high temperatures (e.g., summer), and low relative humidities (e.g.,
the Southwest). Furthermore, if sufficient coarse particles exist that can react with nitric acid
(e.g., sea salt, alkaline soil dust), coarse nitrate particle formation is favored. As subsequent
discussion bears out, these generalizations based on thermodynamic equilibrium explain much
of observed nitrate aerosol behavior. ••••.,. , , ;-,
The extreme volatility of particulate nitrate makes its measurement extremely difficult
and uncertain (Sloane and White, 1986). Significant positive and negative artifacts can occur
with different measurement techniques using different filter media. Glass fiber filters have a
significant positive nitrate particle artifact because they tend to sorb nitric acid vapor (Appel
et al., 1985). Significant particle nitrate loss from quartz filters in storage was noted by
Dunwoody (1986). Teflon filters may have a significant negative artifact because nitrate is
volatilized during or after the sampling, resulting in a potential underestimation of nitrate
particle mass. Volatilization losses of fine particle nitrate are likely to be even greater for a
Teflon filter loaded with acid sulfate than from a clean filter. As reported by a number of
researchers, actual particle nitrate concentrations tend to be about 20 to 50 percent higher
than Teflon-based measurements, on an annual average basis (Appel et al., 1981; Stevens,
1987; John, 1986; Cadle, 1985). However, others (White and Macias, 1987; Malm and
Gebhard, 1988) suggest that at least during winter conditions nitrate particle measurements
might be low by factors of three to five. The denuder/nylon-filter method may be the most
accurate measurement technique (Malm et al., 1989; Allegrini and De Santis, 1989; Stevens
et al., 1988; Mulawa and Cadle, 1985); however, only limited data are available that were
obtained with this technique. Thus, in evaluating empirical studies of the importance of
nitrate to total light extinction, it is important to consider the complications caused by
uncertainty in nitrate particle measurements.
Further complicating the definition of the role of nitrate is the fact that nitrate particles
will absorb water vapor, becoming water solutions, at high humidities (above 62 percent).
The water associated with the nitrate results in scattering efficiencies per unit mass of nitrate
that are much larger than dry particle efficiencies. The effect on light scattering efficiencies
of liquid water associated with aerosols has been known for a long time, but the specific
August 1991
11-18
DRAFT-DO NOT QUOTE OR CITE
-------
1 effect of associated water is difficult to quantify. Empirical studies have used a nonlinear
2 relative humidity term to attempt to account for this effect.
3 . Tang et al. (1981) developed a computer model for calculating the optical properties of
4, nitrate particles, both alone and in combination with sulfate, as a function of particle size and
5 •., relative humidity. This model was based on multicomponent aerosol thermodynamic theory
6 as a function of particle chemical composition and relative humidity. Light scattering
7 ,. efficiencies were calculated from resulting particle sizes using Mie scattering theory.
8 Figure 11^8 through 11-12 summarize the light extinction coefficients for 1 /zg/m3 of sulfate
9 or nitrate aerosol, or both, as a function of humidity. Figure 11-8 shows that pure
10 ammonium sulfate exhibits a deliquescent point at 80 percent relative humidity. At
11 humidities above 80 percent, water vapor condenses, thereby increasing the aerosol particle
12 size, volume, and light scattering. At humidities below 80 percent, the extinction efficiencies
13 ,range from 1 to 4 m2/g of sulfate; while above 80 percent humidity, extinction efficiencies
14 .. can increase considerably above 10 m2/g. Figure 11-9 illustrates the hysteresis effect, that is,
15 , the ability of the particle to hold on to liquid water, that can result when relative humidity is
16 slowly decreased. Figure 1140 shows the increase in light extinction of pure ammonium
17 nitrate aerosol as a function of relative humidity. At and above the deliquescent point at
18 62 percent humidity, the scattering efficiency increases by a factor of two or more because of
19 the condensed water vapor associated with the nitrate particle. Figures 11-11 and 11-12 show
20 the effects of humidity on the light extinction efficiencies of different mixtures of sulfate and
21 . nitrate aerosols. Externally mixed aerosols, those in which the sulfate and nitrate exist on
22 different particles, exhibit the separate deliquescent points for ammonium sulfate (80 percent
23 RH) and ammonium nitrate (62 percent RH). Internally mixed aerosols, in which the sulfate
24 and nitrate occur mixed within the same particle, do not exhibit distinct deliquescent points
25 and have less water associated with them at a given humidity; and hence have lower light
26 extinction efficiencies. The sulfate and nitrate aerosol mixtures may also exhibit hysteresis
27 effects in situations where humidity is reduced, thereby causing a.haze to linger.
28 ,, , . . , -
August 1991
DRAFT-DO NOT QUOTE OR CITE
-------
fc; 10'
UJ
o
El
u.
LU
8
O
10
,-2
10
-3
1.5
2.0
2.5
1.01
50 60 70 80 90
RELATIVE HUMIDITY, %
100
1
2
3
4
5
6
Figure 11-8. Light scattering coefficient for 1 /tg/m3 of sulfate or nitrate aerosol as a
function of relative humidity; bscat versus relative humidity for ammonium
sulfate aerosol (having particle size distributions characterized by D =
0.2 fim and ag = 1.01, 1.5, 2.0, and 2.5). g
Source: Tang et al. (1981). .,
August 1991
11-20 DRAFT-DO NOT QUOTE OR CITE
-------
2.4
2.2 -
0 20
1
ul
i 1.8
o
a i-e
CO
y
g 1-4
Su
1.0
THEORETICAL
EXPERIMENTAL
O
20 30 40 50 60 70
RELATIVE HUMIDITY, %
80
90 100
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
Figure 11-9. Light scattering coefficient for 1 ftg/m3 of sulfate or nitrate aerosol as a
function of relative humidity; bscat versus relative humidity for ammonium
sulfate aerosol showing the effect of hysteresis.
Source: Tang et al. (1981).
11.5 ROLE OF NITROGEN OXIDES IN URBAN HAZE
The most significant effects of nitrogen dioxide and especially nitrate aerosols have been
observed in urban hazes, especially in western urban areas such as those in California; and in
wintertime urban hazes, Such as those in Phoenix and Denver. The importance of nitrate has
been studied empirically by many researchers, but the caveats noted above regarding
measurement uncertainties should be kept in mind when interpreting these studies. In
addition, research has been carried out in recent years on theoretical approaches for analyzing
urban haze chemical reactions and aerosol coagulation.
August 1991
11-21
DRAFT-DO NOT QUOTE OR CITE
-------
-1
10
h-"
UJ
o
E
in
o
o
§ io-2
10
-3
1.01
1.5
2.0
50 60 70 80 90
RELATIVE HUMIDITY, %
100
1
2
3
4
5
6
Figure 11-10. Light scattering coefficient for 1 jtg/m3 of sulfate or nitrate aerosol as a
function of relative humidity; bscat versus relative humidity for
ammonium nitrate aerosol (having particle size distributions characterized
by Dg = 0.6
and ot = 1.01, 1.5, and 2.0).
Source: Tang et al. (1981).
August 1991
11-22 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
10
-1
yj
O
LJL
LL
111
O
O
10'
CASS(1979)
EXTERNAL
MIXTURE-*
WHITE & ROBERTS (1977)
(0.4,1.
(0.29,1.5)
0.
(0
INTERNAL
MIXTURE
.6,1.5) >
i.2,1.5) ^
50 60 70 80 90 100
RELATIVE HUMIDITY, %
Figure 11-11; Light scattering coefficient for 1 jig/m3 of sulfate or nitrate aerosol as a
function of relative humidity; bscat versus relative humidity for externally
and internally mixed sulfate and nitrate aerosols (S:N = 3:1) for
indicated size distributions (Dg, «g).
Source: Tang et al. (1981).
August 1991
11_23 DRAFT-DO NOT QUOTE OR CITE
-------
10-1
LLJ
g
LL
LL.
LU
O
o
10-2
10-3
EXTERNAL
MIXTURE
INTERNAL
MIXTURE
(0.4,1.5)
(0.29,1.5) 11
(Q.W.S)//
(0.2,1.5)
50 60 70 80 90
RELATIVE HUMIDITY, %
100
1
2
3
4
5
6
Figure 11-12. Light scattering coefficient for 1 /tg/m3 of sulfate or nitrate aerosol as a
function of relative humidity; bscat versus relative humidity for externally
and internally mixed sulfate and nitrate aerosols (S:N = 1:2) for
indicated size distributions (D , a ).
<5 O
Source: Tang et al. (1981).
August 1991
11-24 DRAFT-DO NOT QUOTE OR CITE
-------
1 11.5.1 California Urban Areas
2 The most work has been done and the most significant effects 'of NOX have been
3 observed in California, especially in the Los Angeles area. White and Roberts (1977) studied,
4 the statistical relationships between light scattering coefficient and the aerosol constituents of
5 Los Angeles area smog measured during the summer and early fall of 1973 as part of the
6 Aerosol Characterization Experiment (ACHEX). Using linear regression techniques, they
7 estimated that nitrate aerosols contributed, on average, about 27 percent of the total light
8 scattering coefficient. Nitrates were found to have a light scattering efficiency, having units
9 of m2/g of nitrate anion (multiply anion mass by 1.3 to obtain total ammonium nitrate mass),
10 of 1.7 ± 3.9 ju2, where \i is the relative humidity (in percent divided by 100). Thus, at a
11 humidity of 50 percent, the light scattering coefficient of nitrates was estimated to be
12 2.7 m2/g. Appel et al. (1985) have commented that White and Roberts (1977) may have
13 seriously underestimated nitrate scattering efficiencies because the glass fiber filters used to
14 collect aerosol samples had a strong positive artifact (i.e., gaseous HNO3 was deposited on
15 the filter, thereby inflating the nitrate aerosol measurement).
16 Cass (1979) used linear regression to study the relationships between sulfate and nitrate
17 concentrations and visibility in Los Angeles over the decade, 1965 through 1974. Sulfates
18 • and nitrates were found to be significant contributors to total light extinction. The best fits to
19 measured visibility were obtained with regression coefficients of the form, ft I (I - ^), where /*
20 is the relative humidity as defined previously: This is indicative of hygroscopic or
21 deliquescent properties of sulfate and nitrate. The values for ft for sulfate and nitrate were
22 4.1 and 2.5 m2/g, respectively. For a relative humidity of 50 percent, this would yield
23 overall respective light extinction efficiencies for sulfate and nitrate and associated water of
24 8.2 and 5 m2/g.
25 Trijonis et al. (1982) investigated the visibility-aerosol relationship in California using
26 data from 34 locations. They found that NO2 contributed a rather uniform 7 to 11 percent of
27 total light extinction (bext) throughout California. Although they were not of adequate quality
28 to make definitive statements, the data suggest that nitrates are more important contributors to
29 bext in northern California, where they may contribute 10 to 40 percent of bext.
30 In probably the most accurate assessment of the role of nitrate, Appel et al. (1985)
31 studied the relationship between light extinction and aerosol composition using empirical data
August 1991
11-25
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
from three California cities: San Jose, Riverside, and Los Angeles. Multiple linear
regression was used to relate aerosol concentrations to measured light extinction. Emphasis
was placed in the study on minimizing the sampling artifacts that may have caused substantial
errors in previous studies. A linear regression equation that related concentrations of aerosol
species and NO2 to measured bext provided a very good fit (r = 0.96), with average
agreement within 6 percent. The regression coefficient for fine particle nitrate was found to
be higher than in other studies. The coefficient, interpreted as the light extinction efficiency
of nitrate and associated water, was 6.3(1 + /*) m2/g, where /* is the relative humidity as
defined previously. Using the average humidity of 46 percent measured during the sampling
yields a nitrate light extinction efficiency of 9.4 m2/g. The average sulfate extinction
efficiency was comparable: 8.6 m2/g for the average humidity of the sampling period.
Nitrate aerosol, on average, contributed 36 percent of total light extinction in these three
cities; NO2 contributed 8 percent. -
The contributions of various aerosol constituents to total light extinction at two locations
in the Los Angeles area in 1980 were estimated by Pratsinis et al. (1984). Particulate nitrate
aerosol contributed 4 and 1 percent and NO2 contributed 7 and 6 percent of the light
extinction in Lennox and Duarte, respectively. The contributions of the NOX species were
considerably less than those from sulfate (31 to 53 percent), organics (13 to 23 percent), and
elemental carbon (14 to 21 percent). Gray et al. (1984) determined that nitrate aerosol
contributed 13 percent of the light extinction in downtown Los Angeles in 1982.
11.5.2 Urban Areas in the Western United States
Outside of California, the most significant urban hazes that have been shown to be
associated with NOX occurred in the winter in Denver and Phoenix. Nitrogen oxides, both
NO2 and nitrate aerosol, were found to be significant contributors to the winter haze in
Denver (Groblicki et al., 1981; Lewis et al., 1986). Multivariate statistical analysis
(regression) was used to analyze the relationships between light scattering and absorption and
concentrations of particles and gases measured on 41 consecutive days in November and
December 1978. Most of the light extinction was found to be caused by particles < 2.5
in diameter. Elemental carbon (soot) was found to be the most significant contributor,
accounting for 37 percent of light extinction above natural Rayleigh background. Sulfate
August 1991
11-26
DRAFT-DO NOT QUOTE OR CITE
-------
1 (and associated water) was found to contribute 20 percent, nitrate (and associated water)
2 17 percent, and organic carbon 13 percent; the remaining fine particle matter contributed
3 7 percent, and NO2 contributed 6 percent. All measurements were based on a wavelength of
4 light of 0.475 jwm (Hasan and Dzubay, 1983). The contribution of NO2 would have been
5 larger at the blue end of the visible spectrum. If the contribution of nitrate and NO2 are
6 combined, the total NCL contribution to Denver winter haze is 23 percent, second only to the
•"•-•', " L
7 contribution of elemental carbon. Wolff et al. (1981) determined the emission source
8 contributions to the Denver winter haze. Of the total NOX contribution to the winter haze of
9 23 percent, combustion of natural gas, oil, and coal (in power plants and boilers) accounted
10 for more than half (14 percent), while automotive contributions were the largest part of the
11 remainder (9 percent). Hasan and Dzubay (1983) developed estimates of light extinction
12 efficiency of various aerosol components of the 1978 Denver winter haze using both
13 regression analysis and Mie scattering theory based on measured particle size distributions.
r\
14 For ammonium nitrate aerosol, regression gave a scattering efficiency of 2.4 to 2.5 m /g,
r\
15 while theoretical calculations yielded a scattering efficiency of 3.7 to 3.8 m /g.
16 Solomon and Moyers (1984) studied the contributors to light extinction in Phoenix,
17 Arizona,.during January 1983, when winter hazes were observed. Elemental carbon was by
18 far,the largest contributor to light extinction, at 41 percent of bext, on the average.
19 Approximately equal contributions resulted from nitrate (15 percent), organic carbon
20 (15 percent), and sulfate (13 percent). The contribution from NO2 averaged 3.2 percent.
21 Solomon and Moyers (1986) reported that the fine nitrate aerosol measured in Phoenix in
22 January 1983 was 13.4 percent of the total fine particle mass, comparable to the 12.2 percent
23 contribution of nitrate found in Denver during November and December 1978 and much
24 higher than the contribution reported in other major metropolitan and rural areas. They
25 concluded that motor vehicle emissions accounted for most of the nitrate and other fine
26 particle mass that caused the observed haze.
27 Dzubay et al. (1982) studied the relationships between visibility and aerosol composition
28 during summer in Houston, Texas. Nitrate was found mainly on coarse particles and was
29 determined to be an insignificant (0.5 percent) contributor to the total light extinction. It was
30 conjectured that fine nitrate aerosol did not condense because the sulfate was not fully
31 neutralized (i.e., there was insufficient ammonia to react with HNO3); and that HNO3
August 1991
11-27
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
condensed on the alkaline coarse particles, which were a significant sink for nitrate. Nitrate
particle measurement artifacts may also have been a major factor in this study. Nitrogen
dioxide contributed 4.7 percent of bext.
Stevens et al. (1988) reported measurements made during the winter of 1986-1987 in
Boise, Idaho. There nitrate aerosol was a significant component of total light extinction,
contributing 13 percent of the fine particle mass. Less than 10 percent of the total nitrate was
left in the vapor phase as nitric acid. This study represents one of the few good sources of
fine particle nitrate data; measurements in this study were made using an annular denuder
followed by Teflon® and nylon filters.
11.5.3 Urban Areas in the Eastern United States, Europe, and Mexico
Few studies of the role of nitrate aerosol in visibility impairment have been conducted
outside of the western United States. Furthermore, because of the significant problems
inherent in measuring particulate nitrate (discussed in preceding sections), the studies
described here may not have accurately measured particulate nitrate concentrations. This
problem should be kept in mind. Nitrate aerosol contributions appear to be lower in the
eastern United States than in California and other western U.S. cities, perhaps because of
higher sulfate concentrations competing for the available atmospheric ammonia. Using,
multiple linear regression techniques, Trijonis and Yuan (1978a) found that nitrate did not
account for any of the observed light extinction in most of the cities in the northeastern and
north central United States. Nitrates accounted for 8 percent of total light extinction in
Columbus, Ohio. There the light extinction efficiency of nitrate was estimated from
regression analysis to be in the range of 6 to 9 m2/g.
Wolff et al. (1982) found that nitrate contributed minimally to light extinction in Detroit
during July 1981. Fine particle nitrate averaged 0.2 ^g/m3; coarse particle nitrate was
higher, at 1 /*g/m3. This was consistent with other measurements made in the eastern United
States (Ferman et al., 1981), where little nitrate was found in the fine fraction. Nitrogen
dioxide contributed 4 percent of bext in the Wolff et al. study (1982).
Colbeck and Harrison (1984) found significant quantities of nitrate aerosol in northwest
England. Visibility there was strongly correlated with both nitrate and sulfate concentrations.
Diederen et al. (1985) investigated the nature of the haze in western Netherlands during the
August 1991
11-28 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
period 1979 to 1981. Ammonium nitrate aerosol was found to contribute 35-percent of total
bext, and NO2 to contribute 2 percent.
Bravo et al. (1988) found high concentrations of nitrate aerosol and NO2 in Mexico
City (6.4 jug/m3 and 0.07 ppm); however, the contributions of these species to the total light
extinction budget were relatively small (5 and 2.5 percent) because of the much higher
concentrations of other aerosol species. Total light extinction was dominated by soot
(31 percent), sulfate (30 percent), organics (15 percent), and other species (16 percent).
'11.5.4 Modeling Urban Haze Effects
Russell and Cass (1986) developed a Lagrangian trajectory model that incorporates
gaseous and aerosol chemistry and aerosol equilibrium. This model was applied to a smog
episode in southern California. Predictions from the model compared well with
measurements of O3, NO2, HNO3, NH3, PAN, and particulate nitrate. When the model was
used to investigate alternative control techniques for nitrate aerosol, NOX emission control
was found to produce a nearly proportional (linear) reduction in total nitrate (HNO3 vapor
plus particulate nitrate) and slightly greater than proportional reductions in particulate nitrate.
Particulate nitrate concentrations were found to be most effectively reduced by reducing
ammonia (NH3) emissions, especially from farm-related activities.
Russell et al. (1988) developed a grid-based Eulerian airshed model that incorporates a
chemical reaction mechanism for gaseous and aerosol species. The model was compared with
measurements and the model calculations of aerosol nitrate concentrations were found to be in
good agreement with measurements.
11.6 ROLE OF NITROGEN OXIDES IN NONURBAN REGIONAL HAZE
The effects of NOo and nitrate on regional haze outside of urban areas appear to be less
. • ' - , / £•*
significant than their effects on urban hazes. Nitrogen oxides may not be significant in these,
nonurban regional hazes because of low concentrations of nitric acid (HNO3) and ammonia
(NH3), high ambient temperatures, and low humidities in the West; and because of high
sulfate concentrations in the East that compete for available,ammonia.
August 1991
11_29 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
11.6.1 Nonurban Areas of the Western United States
As noted previously (Section 11.5.3), few studies of the role of nitrate aerosol in
visibility impairment have been conducted outside of the western United States, r
Furthermore, because of the significant problems inherent in measuring particulate nitrate
(discussed in preceding sections), the studies described below may not have accurately
measured particulate nitrate. , <
Macias et al. (1981) found that nitrate made small or negligible contributions to regional
haze at one site in Arizona on several monitoring days in the summer and winter of 1979,
although on one day ammonium nitrate was about 8 percent of the fine particle mass.
White and Macias (1987) found very low concentrations of nitrate aerosol in the
nonurban, intermountain West. Measurements of nitrate aerosol concentrations averaged
0.09 ng/m3. Nitrate was very episodic, however, with major contributions to this average
arising from a small number of episodes. Higher concentrations were observed in the North
and at all sites during the winter, White and Macias (1987) commented that during the :
winter the measurements may have underestimated nitrate aerosol concentrations by as much
as a factor of three because of nitrate volatilization from the filters.
Trijonis et al. (1988) analyzed data collected in the Mohave Desert of California over a
2-year period, 1983 to 1985, to determine the species contributing to light extinction. They
found that for both average and worst-case conditions the sum of particulate nitrate and NO2
contributed 13 + 5 percent of bext. .
Malm et al. (1989) evaluated the contribution of nitrate aerosol, along with larger
contributions from sulfate and carbonaceous aerosols, to wintertime visibility impairment in
the scenic Southwest near Grand Canyon and Canyonlands national parks. Nitrate
concentrations during January and February 1987 at Grand Canyon averaged 0.1 to
0.3 jig/m3. Multiple linear regression analysis suggested that nitrate particles had an average
scattering efficiency of 4.7 m2/g and contributed 6 to 14 percent of the fine particle light
extinction during the wintertime study. Nitrate was generally a much smaller contributor,
however, to light extinction than sulfates, which contributed 62 to 72 percent of fine particle
extinction, and organics, which contributed 15 to 16 percent.
August 1991
11-30 DRAFT-DO NOT QUOTE OR CITE
-------
1 11.6.2 Nonurban Areas of the Eastern United States
2 Mathai and Tombach (1987), in their review of visibility and aerosol measurements in
»•>
3 the eastern United States, concluded that fine nitrate concentrations averaged 1 jtig/m . In the
4 studies they summarized, fine-particle nitrate had been measured for very short (week and
^t
5 month) periods and concentrations had ranged from 0.-2 to 0.9 jitg/m .
6 Wolff and Korsog (1989) found that NO2 (averaging 4 ppb) accounted for less than
•7 ,•- 1 percent of total light extinction in the Berkshire Mountains of Massachusetts in the summer
8 of 1984. Sulfate and associated water caused most (77 percent) of the light extinction.
9 Nitrate aerosol was not found. The measurements of Vossler et al. (1989) at Deep Creek
10 Lake in Maryland and of Pierson et al. (1987) in the Allegheny Mountains were consistent
11 . • with the Berkshire Mountains study; NO2 averaged 4 ppb, 'and nitrate aerosol concentrations
12 were very small relative to sulfate. The latter two studies, unlike the Berkshire study, used
13 the more accurate denuder-nylon filter samples.
14 Dzubay and Clubb (1981) found that for summer conditions in Research Triangle Park,
15 North Carolina (nonurban but near urban areas), NO2 light absorption accounted for only
16 2 percent of total light extinction. Particle scattering caused most of the light extinction
17 (75 percent), followed by Rayleigh scattering from air (7 percent) and particle light
1,8' absorption (7 percent).
19 - ,- - .- - • ' ••• '- • ' •'' -
20 11.6.3 Modeling Regional Haze Effects
21 Latimer et al. (1985a) used a Lagrangian regional visibility model and emission
22: : inventories for the southwestern United States to estimate the effects of manmade emission
23 sources on regional visibility in 1980 and 1995. In this assessment, nitrate aerosol was found
24 to be a potentially significant contributor to the manmade portion of nonurban regional haze.
25 While manmade sulfate sources were found to be the largest contributors to haze, contributing
26 over half (50 to 60 percent) of the manmade fraction, nitrate was estimated to be the next
27 "•' largest contributor (10 to 20 percent). Although manmade organic and elemental carbon
28 contributions to regional haze were found to be small (less than 10 percent of the manmade
29 fraction), biogenic organic aerosol was estimated to be a large contributor to total light
30 extinction (the sum of natural and manmade fractions).
. August 1991
11-31
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
In this modeling study, it was cautioned that the estimates of the contribution of nitrate
to the manmade total were uncertain because of uncertainties in the relative distribution of the
nitrate anion (NO^) between the optically inactive nitric acid (HNO3) and the light-scattering
ammonium nitrate aerosol (NH4NO3). This uncertainty resulted largely from the uncertainty
regarding background concentrations of ammonia (NH3), which is essential to the formation
of ammonium nitrate aerosol. On the basis of thermodynamic equilibrium considerations, the
study showed that nitrate aerosol would be most likely to condense in winter and least likely
in summer. Nitrate aerosol was found to be a significant portion of increases in regional
haze projected for the period 1980 to 1995. Latimer et al. (1985b, 1986) evaluated the
performance of this regional visibility model by comparing model calculations with
particulate, visibility, and wet deposition measurements performed by the U.S. Environmental
Protection Agency, the National Park Service, and the Electric Power Research Institute.
This comparison showed that model predictions of sulfate and nitrate concentrations and light
extinction were relatively unbiased and were highly correlated with actual measurements.
The average nitrate aerosol concentration predicted by the model was 0.22 /*g/m3,
approximately 2.4 times the average measured during the Western Regional Air Quality Study
in 1981 of < 0.1 /ig/m3 that was reported in Tombach et al. (1987) and the'value of
0.09 Atg/m3 reported by White and Macias (1987).
« ' r
Latimer et al. (1986) and Latimer (1988c) applied this regional visibility model to the
case of winter layered haze observed near the national parks in Utah and Arizona. An
average nitrate aerosol concentration of 0.35 /wg/rn3 was predicted. This value compares
reasonably well with the average of 0.16 jug/m3 measured during a special study in 1986
(Latimer, i988c) and the average of 0.38 ^g/m3 measured during the WHITEX experiment
in 1987 (Malm and Iyer, 1988). However, the model underpredicted the observed sulfate
concentrations by a factor of two to four. Although considerable uncertainty exists over the
accuracy of nitrate measurements (Malm and Gebhard, 1988), nitrate may be a significant
contributor to winter layered haze (approximately 15 to 25 percent of extinction from
manmade sources (Malm et al., 1989), even though sulfate appears to be the dominant
contributor,
Latimer (1988a) developed a spreadsheet template for calculating the effect of changes
in aerosol species concentration on total light extinction and visibility. As part of that efforfj
August 1991
11-32
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13.
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
available measurements of particle chemical composition and concentration and visibility or
light extinction were compiled. Using an assumed nitrate light-scattering efficiency of
8 m2/g, Latimer (1988a) estimated the relative contribution of nitrate to total light extinction
in numerous locations where both aerosol and visibility data were available. Nitrate generally
contributed less than 10 percent to total extinction, except in Portland, Oregon, where it was
11 to 14 percent; Denver, Colorado, 16 percent; Los Angeles, California, 20 percent; and
Riverside, California, 40 percent. Latimer (1988a) found that measured visual ranges agreed
well with visual ranges derived from the measured aerosol constituents and their respective
light extinction efficiencies.
11.7 ROLE OF NITROGEN OXIDES IN PLUME VISUAL IMPACT
Much of the regulatory attention that has been given to visibility during the past decade
has focused on the issue of the visibility impacts of plumes from individual emission sources.
This plume visual impact is commonly called "plume blight" (U.S. Environmental Protection
Agency, 1979). Particularly in areas of pristine background visibility, such as the
intermountain West, the visual impact of plumes such as those from power,plants can be
quite significant as far as 100 km from sources (U.S. Environmental Protection Agency,
1979; Latimer, 1979, 1980). Considerable work has been carried out during the past decade
to develop and evaluate plume visual impact computer models and to develop technical
guidance for plume visual impact evaluation as part of the implementation of U.S. EPA's
visibility regulations under the visibility protection provisions of the Clean Air Act. Nitrogen
dioxide has been found to be a significant contributor to plume visual impact from modern,
well-controlled power plants.
The contrast of a plume against an optically thick horizon sky background can be
calculated by solving Equation 11-1 .(Latimer et al., 1978;'White et al., 1986):
- exP(-Tplume)] [exp(-bext rp)],
(11-8)
August 1991
11-33
DRAFT-DO NOT QUOTE OR CITE
-------
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
where
cplume = contrast of the plume against the horizon sky
J = source function defined previouly;
rplume = optical thickness of the plume (Jbextdr); '
^ext = extinction coefficient of the intervening background atmosphere between the
plume and the observer; and
rp = distance between the plume and the observer. ?
For a pure NO2 plume, the first term (in the first pair of square brackets) equals -1, and
therefore Cplume is always negative, signifying a dark plume. If one also assumes either that
the plume is very close to the observer (rp « 0) or that the intervening atmosphere is optically
thin (bext« 0), then the last term in this equation equals 1, and the following equation for an
NO2 plume is obtained:
Cplume = -[
= ~tt ~ exPCf
plume bag
(11-9)
If one assumes that Cplume must equal at least -0.02 for a plume to be visible, then the
plume optical thickness (rplume) must be at least 0.02. For a plume that is 1 km wide, this
optical depth can be caused by 0.065 ppm (122 /*g/m3) of NO2 at A = 0.55 ^m or by
0.012 ppm (22 jig/m3) at A = 0.4 pm. For a plume 10 km wide, the same effect could be
caused by NO2 concentrations one-tenth as large. Melo and Stevens (1981) found that under
typical conditions a plume NO2 optical thickness corresponding to 90 ppm-km (or 0.090 ppm
in a 1 km wide plume) was required to make a plume just visible against a blue horizon sky
background. Using a predecessor of the PLUVUE models (Johnson et al., 1980; Seigneur
et al., 1984), Latimer (1980) investigated the relationship between NOX emission rates from
power plants and plume contrast and other optical parameters. He found that the yellow-
brown coloration of the power plant plume was dominated by NO2 for the modeled cases.
Melo and Stevens (1981) confirmed the dominant importance of NO2 to coloration in an
actual power plant plume. Latimer (1979, 1980) modeled the visual impacts of power plants
of various sizes and NOX emission rates and concluded that yellow-browa plumes could be
observed as far as 100 to 150 km away from a power plant, but only on a few days per year.
White and Patterson (1981) developed nomographs that allow one to determine the
optical properties and relative importance of emitted particles and NO2 as a function of the
August 1991 H_34 DRAFT-DO NOT QUOTE OR CITE
-------
1 scattering angle and the particle size distribution. Vanderpol and Humbert (1981) identified
2 NO2 as the primary plume colorant when particle size was greater than 0.5 jim. Haas and
3 Fabrick (1981) performed a sensitivity analysis to investigate the effects of NO2 and particles
4 in plumes on various indicators of color and contrast.
5 In studies of a power plant plume in the southwestern United States as part of the
6 VISTTA project, Richards et al. (1981) never found paniculate nitrate even though nitric acid
7 vapor was formed at rates 3 to 10 times the rate at which sulfate aerosol was formed. They
8 concluded that nitrate aerosol did not condense because of inadequate background
9 concentrations of ammonium ion. Hegg and Hobbs (1983) measured the constituents of
10 another power-plant plume in the Southwest and found rapid formation of both nitric acid and
11 nitrate aerosol. Nitrate aerosol constituted 15 to 75 percent of the nitrate in the plume.
12 Measured plume aerosol size was primarily in the 0.25 ^m range. Approximately equal
13 contributions to plume light extinction were made by particles and NO2. The reason the
14 Hegg and Hobbs (1983) findings were quite different from those of Richards et al. (1981) is
15 not clear, but the findings may have differed because background ammonia concentrations
16 differed at the respective sites.
17 Also as part of the VISTTA study, Blumenthal et al. (1981) measured the dispersion,
18 chemistry, and optical properties of a coal-fired power plant plume in the Southwest. On the
19 basis of this measurement program, they concluded that NO2 was the primary plume
20 colorant, that secondary aerosol formation could be neglected within 100 km of the source,
21 and that the PLUVUE model adequately characterized observed effects. Bergstrom et al.
22 (1981) evaluated the PLUVUE model using VISTTA data and found that the model
23 performed reasonably well, but that it slightly overpredicted observed plume visual impacts.
24 Sensitivity analyses performed indicated that NO2 was the principal plume colorant.
25 = The most detailed evaluation of plume visibility models was carried out as part of the
26 VISTTA study (White et al., 1985, 1986). Four plume visibility models, including the two
27 versions of PLUVUE (Latimer and Samuelsen, 1975, 1978; Latimer et al., 1978; Johnson
28 et al., 1980; Seigneur et al., 1984), the ERT visibility model (Drivas et al., 1980),
29 PHOENIX (Eltgroth, 1982), and the Los Alamos visibility model (Williams et al. 1980;
30 1981), were evaluated by comparison with field measurements of plume concentrations,
31 optical parameters, and observed plume color and contrast made at a large power plant in the
August 1991
11-35
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
Southwest, well-controlled for paniculate, and at less well-controlled power plants in the
Midwest and an uncontrolled smelter in the'Southwest. Of the four, the first two, the
PLUVUE and ERT models, were found to be most accurate in predicting the plume visual
impacts observed in the field measurement programs. The plume contrast for the power plant
with modern particulate controls could be adequately explained accounting just for the plume
NO2 concentrations; participates did not play a significant role. In the study of strong
particulate emission sources (White et al., 1986), the performance of PLUVUE II and the,
ERT models was less satisfactory than for the NO2-dominated plumes. However, the
relatively poor performance of these two models may have resulted in large part from the
imprecise specification of model inputs (particle size and background sky radiance). Model
performance was found to depend strongly on model input specification.
11.8 CONTRIBUTIONS OF NITROGEN OXIDES TO THE LIGHT
EXTINCTION BUDGET
Trijonis (1987) evaluated the relative contributions of NO2 and nitrate aerosol to
visibility by comparing the light extinction caused by these NOX species to total light
extinction for field studies in which the total light extinction and the chemical composition of
the haze were measured simultaneously. Tables 11-2 and 11-3 summarize the results for ,
NO2 and nitrate aerosol, respectively. The contribution of NO2 to light absorption was
obtained by multiplying the NO2 concentration by its light absorption efficiency at 0.55 jum
(0.33 Mm"1 ppb"1). The relative contribution of NO2 ranged from < 1 percent of total light
extinction in the rural East to as high as 9 percent in urban areas. Nitrate aerosols generally
contributed less than 10 percent, except in California urban areas and in Denver during the
winter, where the fraction of total light extinction attributable to nitrate was as high as
34 percent. It should be noted, however, that these estimates may be lower bounds because
of nitrate measurement problems resulting frpm the volatility of sampled nitrates.
Trijonis (1987) summarized his assessment of the contribution of the two NOX species
as shown in Table 11-4.
These estimates may be somewhat high for the rural situations if nitrate aerosol
concentrations are limited because of low nitric acid and ammonia concentrations,
August 1991
11-36
DRAFT-DO NOT QUOTE OR CITE
-------
[
F3
I
O
Q
£
.lo
ge N02
tration,
Av
Con
s
pq
00
oo
co
O O O
cs o\ cs
ON O\ >/2 O\
1-1 cs •* cs
00
PQ
August 1991
11-37
DRAFT-DO NOT QUOTE OR CITE
-------
•<* in
CO CO
CN o
a a
co co
en
en
CO CO
CO
00
o\
00
.g
g
w *i
«-« *c?
1
1
£
8
o\
8
SJ
SAROADDa
See footnote <
August 1991
11-38 DRAFT-DO NOT QUOTE OR CITE
-------
z
0
g
Ej
n
H
§
H
o
H
^^
rp
H
g
g
p^^
t^
1
S
u
TABLE 11-3. NITRA
a
o «
311-3
£| a s
w 8 <;
&- 1,
§ o "IJ Jj
**d t*^
1 1
LJ_j +•* en
|0"|-|
^ £H
^i O
< o
1
p
i
T^-U^I— i en >— I*— i i— i •<* l> o\
en eN A en <-i <-i ^ ,-( «
A
oo oo oooo [& ^ •-..'•
ooo in o o oo o o
o\ oo 1-1 o\ o --i in . o\ *-< o
fsl ^H (N^ -^H ^H ^H T*H i~H T~1
oo oo oo oo ^: ^ ' .
^^ in oo
S °°- « * * ^ S ^ ^ M A ^ «>
S^A en-HA5 2 cs o
°1'S ^^ '• ' o -2
1 8- If 1 ? '^ ; 8 w ' § | .
i M x • <3 2 wod ^ ^
M^'3 oo S S oi ON S ^-S-
.a . o\ a B*4-5 oo'-i 4s s ' -3 •
to CJ '•^ 'E3 S«N O\ 4-1 -"^ ^2
»rt . i w 4-i w ^ ^« i-y-w
^^r§ S O^*4"1 !> ',g 0 Ctf" § ^
• "!• ' J * J g |l 1 • -a 13 j g f
t 5 1* -9^1^ -g : g -8,1 ^ |
S * ® o" g -°* fe 00 S3. 6 „ 0 fl, «
- 5- f £i g .5 .f &a IP -fS § &
<3j5 ^ Hopsi.d "S, 4-.°^ QIS ,-.SH
1 -3" • 1 111- 1 — 2 1 s s- § -g,,
i^i - 4^ § ***• <3 2 ^ §a>Op3^ g >a3-
- « rf S SP ^(B-S^ !=* ^-S^^'Si
ca 43 >.O8 43 ,PW^S oooo ^jlh-54? O49 QX co «fc! p W'fc
0 : < . '-.. .' Q- ... 0 . Q .
I_)^H c-) enooi-i cN en PI-I eN PM - ^H pqrn
-------
D
2i
H
L(
73
^%
3
3
H
O
H
B
CD
§
13
P
1
u
S
PJ
o
CO
•A
•j<
3
H
o c5*
*— 3 tO >y
gig*
w|
'~i-i ^ «-•
^t K
fji
C5 o
J5 *.g
4) M i-O
CO •§ O«
-1 °
1
Q
i
.2
S
o\
o
o\
d
.S
f --,j »•-*
§ i
^ u2
' 0
1 i
•fi
||
""I* fl
TO o
^J **H
W > 3
O _j °*
S w1^
@ w i
PH 3 "
^ 11
g is
C^ r4
<
o\
CO
o
t— (
co
"1
OJ
CO
.s
"S
1
J
>— 4
ti_J
^
•4-t
a
u
1
.g
to
i
1
c<-i o
w -a
t|
I 1
PH «S
^ 8
co
CO
°* ~;
»-t CO
*-^ oo
. 0\
^3 S3
o s
>-, 5
CN
ID
CO
o\
CO
°5
o
CO
S
i-<
i
1
!
a
J
M
j
61)
.a
"On
1
1-<
n
t
|
to
£d
S "°
d- S
•3 |
-1
J «
O *"
1^ CO
CO
•* cs
A A
o vi
CO i-l
d d
A A
t
o\
&
T-H
*
1 ^
'^ *a
o S
•S 'a
b 1-1
S ^
o\ la
ta J3 t..
cs
^ A
CO •
0 0
CJ\ O\
t— < t— <
•*
^ d
^ A
to
CO
pi j 8
' "1 CO
55 °
•^ $ d
H S oo
3 o o\
Pn o ^
i i 1
W CD1
^ I S" s i! 1
S3 Si " 5
^ 1 ° 1 § ^
§ ^ !^ 3 1
£ ft "«• 63 J
[T"1 (4-1 CO
F— 1 O ^
S m «S O1
j^i i-Lj
§ ^
?N 05 *5 r- 1 H (D
§ & 5 ii j§ 1
E | §^ i 1 «
A
oo
o
•a
CO
cs
d
A
CO
^
^
1
P
ta
-1
1
t
FLON FILTERS
m
P
S
1
j<
CO
A
o
: »— i
^
d
A
' 1
'is
J
1
*§
1
.Ox
,"«
-4~*
S
-§
|
° ^
S3 ®
^^
||
§ 2 ^1 J .a 15- <§ 5 2 ^
Q O to f^ ^5
CD ^ -^
O
S $
§
-------
O f
H -Z
00
* &-a ^,
W £ I co
•S
August 1991
11-41
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
TABLE 11-4. PERCENTAGE CONTRIBUTION OF
NITROGEN OXIDES TO TOTAL LIGHT EXTINCTION
Location
Metropolitan West
Rural West
Metropolitan East
Rural East
Contribution to Extinction, %
NO2
8 ± 1.5
4 ±2
5 + 1.5
2 + 1.5
Nitrate
15 ± 4
6 ±2 .
5 ± 2
5 ±2 .
Total NOX
> A.
23 ± 4.5 , .
10 + 3 ,
10 ± 2.5
, 7 + 2.5
competition from sulfate for ammonia, and temperatures and humidities that favor the
retention of nitrate in the gas phase.
11.9 SUMMARY OF EFFECTS ON VISIBILITY
Emissions of NOX can contribute significantly to visibility impairment in the form of
plumes and hazes. Nitrogen dioxide (NO^ and. ammonium nitrate aerosol (NH4NO3) are the
optically active species of NOX. Other species, including nitric oxide (NO) and nitric acid
(HNO3), are gases with insignificant optical effects. Nitrogen dioxide is a gas that
preferentially absorbs blue light, thus tending to cause yellow-brown atmospheric
discoloration. There is agreement among many studies that NO2 is a strong and consistent
colorant. Aerosols, however, including nitrate, can cause atmospheric discoloration,
particularly when bright objects are observed or the sun is behind the observer.
Nitrogen dioxide has been shown to be the most significant plume colorant for the
yellow-brown power plant plumes that have been observed, primarily in the western United
States, and that are of current regulatory concern to EPA and the States.
Nitrogen dioxide and nitrate aerosol are significant contributors to urban haze,
especially in California and the western United States. Their combined share of total
extinction can be 20 to 40 percent of total light extinction in such urban areas. In nonurban
areas, NOX appears to be a relatively small contributor to light extinction because NO2,
nitrate aerosol, and ammonia concentrations tend to be lower or because moderate or high
temperatures tend to prevent nitrate aerosol from condensing. Nitrate aerosol does not appear
August 1991
11-42
DRAFT-DO NOT QUOTE OR CITE
-------
1 in areas of high concentrations of sirlfate, such as the eastern United States, mainly because
2 acidic sulfate compounds consume the available atmospheric ammonia that is needed to
3 condense nitrate aerosol from nitric acid vapor.
.4 , Theoretical models have been developed for describing the chemical reactions that result
5 in the formation of optically active NOX species, aerosol dynamics of nitrate aerosol,
6 chemical equilibrium of nitrate-water aerosols, the light scattering and absorption properties
7 as a function of the wavelength of light, and effects on visual range, haze contrasts, and
8 atmospheric color. The available comparison of plume visibility models suggests that the
9 effects of plume NO2 can be accurately predicted but that model predictions of the effects of
10 particles are less adequate. Limited work has been done to develop and test models for
11 urban, layered, and regional haze; but much more work is clearly needed.
12 Measurement of nitrate aerosol is complicated by its volatility. However, newer
13 measurement techniques based on the use of denuders have provided reliable measurements.
14 Because older techniques (such as Teflon filters) can seriously underestimate nitrate aerosol
15 concentrations, care must be taken when interpreting data.
16 : Work is needed to understand the apparently nonlinear effects of NOX emission controls
17 on nitrate aerosol concentrations and resulting visibility effects. Also, work is needed to
18 understand the effects of SO2 emission controls on nitrate aerosol production, because the
19 large-scale reduction of sulfate, which competes with nitrate for available ammonia, may
20 result in increases in nitrate aerosol.
21
22
23 11.10 ECONOMIC VALUATION OF EFFECTS ON VISIBILITY FROM
24 NITROGEN OXIDES
25 The primary effects of NOX on visibility were described in previous sections of this
26 chapter and are believed to be: (1) discoloration, producing a brownish color seen in plumes,
27 layered hazes, and uniform hazes, and (2) reductions in visual range (increases in light
28 extinction), especially in urban areas in the western United States. This section discusses the
29 available economic evidence concerning the value of preventing or reducing these types of
30 effects on visibility. Economic studies have not focused specifically on NOX associated
31 changes in visibility for the most part, but some studies have considered the types of visibility
August 1991
11-43
DRAFT-DO NOT QUOTE OR CITE
-------
1 effects that are associated with NOX. The following summary of economic estimation
2 methods and available results is brief. For more detail see Chestnut and Rowe (1989),
3 Mitchell and Carson (1989), Hschhoff and Furby (1988), Cummings et al. (1986), and Rowe
4 and Chestnut (1982).
5
6 11.10.1 Basic Concepts of Economic Valuation
7 Visibility has value to individual economic agents primarily through its impact upon
8 activities of consumers and producers. Studies of the economic impact of visibility
9 degradation by air pollution have focused on consumer activities. Some commercial
10 activities, such as airport operations, may be affected by visibility degradation by air
11 pollution, but available evidence suggests that the economic magnitude of NOX effects on
12 commercial operations probably is very small. In a 1985 report, the U.S. Environmental
13 Protection Agency concluded that only a small percentage of the visibility impairment
14 incidents sufficient to affect air traffic activity can be attributed primarily to manmade air
15 pollutants (2% to 12% in summer in the eastern United States); and according to the
16 information presented previously in this chapter, NOX would not be expected to be a
17 significant contributor to these incidents. Most economic studies of the effects of air
18 pollution on visibility have, therefore, focused on the aesthetic effects to the individual.
19 It is well established that people notice those changes in visibility conditions that are
20 significant enough to be perceptible to the human observer, and that visibility conditions
21 affect the well-being of individuals. This has been verified in scenic and visual air quality
22 rating studies (Middleton et al., 1983; Latimer et al., 1981; Daniel and Hill, 1987), through
23 the observation that individuals spend less time at scenic vistas on days with lower visibility
24 (MacFarland et al., 1983), and through use of attitudinal surveys (Ross et al., 1987). The
25 intent of visibility-related economic studies has been to put a dollar value on changes in well-
26 being associated with visibility degradation.
27 Welfare economics defines a dollar measure of the change in individual well-being
28 (referred to as utility) that results from a change in the quality of any public good, such as
29 visibility, as the change in income that would cause the same change in well-being as that •
30 caused by the change in the quality of the public good. One way of defining this measure of
31 value is to determine the maximum amount the individual would be willing to pay to obtain
August 1991
11-44
DRAFT-DO NOT QUOTE OR CITE
-------
1 improvements or prevent degradation in the public good (see Freeman, 1979, for more
2 detail). For most goods and services traded in markets, this measure can be derived from
3 analysis of market transactions. For non-market goods, such as visibility, this economic
4 measure of value must be derived some other way.
5 For purposes of this discussion, consumer values for changes in visibility can be divided
6 into use and non-use values (there are slight variations in the way these are defined by
7 different economists). Use values are related to the direct influence of visibility on the
8 current and expected future activities of an individual at a site. Non-use values are the values
9 an individual places on protecting visibility for use by others (bequest value) and on knowing
10 that it is being protected regardless of current or future use (existence value). Total value,
11 combining use and non-use, is sometimes called preservation value.
12 . ' "
13 11.10.2 Economic Valuation Methods for Visibility
14 Two main economic valuation methods have been used to estimate dollar values for
15 changes in visibility conditions in various settings: (1) the contingent valuation method, and
16 (2) the hedonic property value method. Both methods have important limitations, and
17 uncertainties surround the accuracy of available results for visibility. Ongoing research
18 continues to address important methodological issues, but at this time some fundamental
19 questions remain unresolved (Chestnut and Rowe, 1989; Mitchell and Carson, 1989;
20 Fischhoff and Furby, 1988; Cummings et al., 1986). Recognizing these uncertainties is
21 important; but the body of evidence as a whole suggests that economic values for changes in
22 visibility conditions are probably substantial in many cases and that a sense of the likely
23 magnitude of these values can be derived in some instances from the available results
24 (Chestnut and Rowe, 1989).
25
26 11.10.2.1 Contingent Valuation Method
27 The contingent valuation method (CVM) involves the use of surveys to elicit values that
28 respondents place on changes in visibility conditions (see Rowe and Chestnut, 1982; Mitchell
29 and Carson, 1989; Cummings et al., 1986; for more details on this method). The most
30 common variation of the CVM relies on questions that directly ask respondents to estimate
31 their maximum willingness to pay (WTP) to obtain or prevent various changes in visibility
August 1991
11-45
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
conditions. The potential changes in visibility conditions are usually presented to the
respondents by means of photographs and verbal descriptions; and some hypothetical payment
mechanism, such as a general price increase or a utility bill increase, is posed.
The CVM offers economists the greatest flexibility and potential for estimating use and
non-use values for visibility. There are many types of changes in visibility for which total
values cannot be derived from market data. As a result, most recent visibility value
applications use the CVM. This approach continues to be controversial, however, and there
are those who question whether the results are useful for policy analysis (Fischhoff and
Furby, 1988). Cummings et al. (1986) and Mitchell and Carson (1989) have conducted the
most comprehensive reviews of the CVM approach to date and have concluded that there is
sufficient evidence to support the careful use of results from well-designed CVM studies in
certain applications.
Among the fundamental issues concerning the application of CVM for estimating
visibility values are the ability of researchers to present visibility conditions in a manner
relevant to respondents and to design instruments that can elicit unbiased values; and the
ability of respondents to formulate and report values with acceptable accuracy. As with any
survey instrument, it is important that the presentation be credible, realistic, and as simple as
possible. The optimal level of detail and the most critical pieces of information necessary in
the presentation to respondents to obtain useful CVM responses continues to be a topic of
research and discussion. Another important issue in CVM visibility research concerns the
ability of respondents to isolate values related to visibility aesthetics from other potential
benefits of air pollution control such as protection of human health. Preliminary results
(Irwin et al., 1990; Carson et al., 1990) suggest that simply telling respondents before asking
the WTP questions to include only visibility is not adequate and may cause some upward bias
in the responses.
11.10.2.2 Hedonic Property Value Method
The hedonic property value method uses relationships between property values and air
quality conditions to infer values for differences in air quality (see Rowe and Chestnut, 1982;
and Trijonis et al., 1984; for more detail on this method). The approach is used to determine
the implicit, or "hedonic," price for air quality in a residential housing market based on the
August 1991
11-46
DRAFT-DO NOT QUOTE OR CITE
-------
1 theoretical expectation that differences in property values that are associated with differences
2 in air quality will reveal how much households are willing to pay for different levels of air
3 quality in the areas where they live. The major strength of this approach is that it is that it
4 uses real market data that reflect what people actually pay to obtain, improvements in air
5 quality in association with the purchase of their homes. The method can provide estimates of
6 use value, but non-use values cannot be estimated with this method..
7 . There are many theoretical and empirical difficulties in applying the hedonic property
8 value method for estimating values for changes in visibility, but the most important limitation
9 is the difficulty in isolating values for visibility from other effects of air pollution at the
10 residence. Hedonic property value studies to date provide estimates of total value for all
11 perceived impacts resulting from air pollution at the residence, including health, visibility,
12 soiling, and damage to materials and vegetation. The potential for estimating separate values
13 for visibility with this method is limited for two reasons. First, the actual effects of air
14 pollution often are highly correlated, making it. difficult to separate them statistically using
15 objective measures. Second, individuals are likely to perceive a correlation between these
16 effects and to act accordingly in their housing decisions, even if the effects are actually
17 separable using objective measures. ,
18 . ' , '
19 11.10.3 Studies of Economic Valuation of Visibility
20 Economic studies have estimated values for two types of visibility effects potentially
21 related to NOX: (1) use and non-use'values for preventing the types of plumes caused by
22 power plant emissions, visible from recreation areas in the southwestern' United States; and
23 (2) use values of local residents for reducing or preventing increases in urban hazes in several
24 different locations.
25
26 11.10.3.1 Economic Valuation Studies for Air Pollution Plumes
27 Three CVM studies have estimated on-site use values for preventing an air pollution
28 , .. plume visible from recreation areas in the southwestern United States (Table 11-5). One of
29, these studies (Schulze et al., 1983) also estimated total preservation (use and non-use) values
30 held by visitors and non-visitors for preventing a plume at the Grand Canyon. A fourth study
August 1991
11-47
DRAFT-DO NOT QUOTE OR CITE
-------
•w-
Valuatio
method
of value
o 5
Study Subject
CO
«
&
•a
§
Jl'
Sei
'
S 3
O
CO
o\
1 .1
8 S
•0 >
a
§
en
CO
o\
73
•s
i
a
8
"
rt 2
m
B
p<
|I I
•J3 o! Oj
a P K
o
GO
0\
i
g
-0
'S S3 . - i*
S
|fi|l|ll
o.g ^D^-?-?
°^ -a
N S
•to- a.
•
I
§
-«
Is
•5 '-S ,
fl ^
-
'S ft'S o
o ts •? e
§
ON
ark visitors
PH
0)
•g
LLJ en
IS
' c Q r
l
'ca H
Q ts
t>
OS
•
I-
73
•4->
-------
1 concerning a plume at Mesa Verde National Park (Rae, 1983) was not included because of
i ...
2 methodological problems with the contingent ranking approach used (Ruud, 1987). The
3 plumes in all three studies were illustrated with actual or simulated photographs showing a
4 dark, thin plume across the sky above scenic landscape features, but specific measures such
5 as contrast and thickness of the plume were not reported. Respondents were told that the
6 source of the plume was a power plant or an unspecified air pollution source. In one study
7 (Brookshire et al., 1976), a power plant was visible in the photographs.
8 The estimated on-site use values for the prevention or elimination of the plume ranged
9 from about $3 to $6 (1989 dollars) per day per visitor-party at the park. These value
10 estimates are comparable to values obtained in these and other studies for preventing fairly
11 significant reductions in visual range caused by haze at parks and recreation areas in the
12 Southwest. A potential problem common to all of these studies is the use of daily entrance
13 fees as a payment vehicle. Respondents may, have anchored on the then-typical $2 per day
14 fee and stated an acceptable proportional increase in entrance fees rather than reporting a
15 maximum WTP. This may have caused some downward bias in the responses, but empirical
16 exploration of this question is needed. An alternative payment vehicle to consider might be
17 total expenditures for the trip to the park.
18 The results of the Schulze et al. (1983) study suggest that on-site use values may be
19 easily dwarfed by total preservation values held by the entire population. For example, with
20 average annual visitation at the Grand Canyon of about 1.3 million visitor-parties (about
21 3 people per party), annual on-site use values for preventing a visible plume every day would
22 be about $8 million based on the Schulze et al. results, while the implied preservation value
23 for preventing a visible plume most days' (the exact frequency was not specified) at the Grand
24 Canyon would be about $5.7 billion each year when applied to the total United States
25 population. There is, however, considerable uncertainty in the preservation value estimates
26 from this study. Chestnut and Rowe (1990) found that the Schulze et al. (1983) estimates for
27 haze at national parks in the Southwest are probably overstated by a factor of two or three
28 and the same probably applies to the plume estimates.
29
30 •.;'- '.
August 1991'
11-49
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
11.10.3.2 Economic Valuation Studies for Urban Haze
Six economic studies concerning urban haze caused by air pollution are summarized in
Table 11-6. Five of these are CVM studies and one is a hedonic property value study.
Although many other hedonic property value studies concerning air quality have been
conducted (see Trijonis et al., 1984; and Rowe and Chestnut, 1982; for reviews), this
California study (Trijonis et al., 1984) is the only one that has used visibility as the measure
of air quality.
The magnitudes of the changes in visual range considered in each study vary, making .
direct comparisons of the results difficult. In Table 11-6 implicit values obtained for a
10 percent change in visual range are reported to allow a comparison of results across the
studies. These estimates are based on an assumption that economic values are proportional to
the percentage change in visual range. All of these studies relied on not. exactly random, but
more or less representative samples of residents in each of the selected urban areas. A range
of socioeconomic characteristics and of neighborhood pollution levels was included in each
sample.
The first five studies in Table 11-6 all focused on changes in urban hazes with fairly
uniform features that can be described as changes in visual range. The sixth study (Irwin
et al., 1990) focused on visual air quality in Denver, where a distinct edge to the haze is
often noticeable, making visual range a less useful descriptive measure because it would vary
depending on the viewpoint of the individual and whether the target was in or above the haze
layer. The studies conducted in Denver and in the California cities are the most relevant '
because hazes in these cities are likely to have a higher NOX component than in the eastern.
cities; but none of these studies focused specifically on NOX.
The California studies in Los Angeles and San Francisco provide some interesting
comparisons because two different estimation techniques were applied for the same locations.
Property value results for both cities were found to be higher than comparable values
obtained in the CVM studies, although Graves et al. (1988) have reported that subsequent
analysis of the property value data revealed that the WTP estimates are more variable than
the original results suggest. The higher property value results are expected because those
results are likely to reflect concern about health as well as aesthetic effects of air pollution.
August 1991
11-50 DRAFT-DO NOT QUOTE OR CITE
-------
O
o
oo
B
00
o5
60
ON
OO
00
g
I
1
CO
1
1
CM
CM
T-H
•6O-
Light extinction based
on airport .visibility
data
1
&
_y
O D
ffl ^
ON
OO
S
w
•i
60
! "
*' 3 ' ' '
1
1
1
T— 1
«*
3
OO
CO
&!!
•§ £^-!3
.•I|l! -
&|S1*
• -d 3 rt — S :
M '« 5 is a
G.'P £ £ &
1
"j*j CO 5? Ctf
^ S S te
.j^ t|"^ .juj p^
J J
1 s
1 ^
1 §fl »
ll
OH
U "Q
CM
oo
Q *^H
o - •> - ^ S
1 w> ^ a ,2 -9 „ 'd" g
.| J | ^ 1 U J J
T-H
1 ^
U ' -w
o
s >^»
§ 0 0
ca .s eo «j
•S'^H
a) .-is s?l S ^
w>r3 O -1 -&
g WD rt CO ^
_S 'w *^ t-i
O 'P '^ <8 a*
1
8ti
^S s S
1 .a, § 1
fl ^ *-« oo
O VH S^ O
P^ >l*^ ,Q ^^
o o
'is '«
•1 §
g CT1
*i ^
8 E
60 &
9 '
'^ "S
rt o
O H
' U -3
CM
oo
ON
i— (
^
ri
1
oo
ON
1— 1
(U
August 1991
11-51
DRAFT-DO NOT QUOTE OR CITE
-------
August 1991
11-52 DRAFT-DO NOT QUOTE OR CITE
-------
1 Both of the CVM studies in California asked respondents to consider health and visual effects
2 but used different techniques to have respondents partition the total values. They found that,
3 on average, respondents attributed about one-third to one-half of their total values to aesthetic
4 visual effects. In spite of many similarities in the approaches used, the CVM results for San
5 Francisco are notably higher than for Los Angeles when adjusted to a comparable percentage
6 change in visual range. One potentially important difference in the presentations was that
7 Loehman et al. (1981) defined the change in visibility as a change in a frequency, distribution
8 rather than simply a change in average conditions. This type of presentation is more realistic
9 but more complex; and it is unclear how it may affect responses relative to presentation of a
10 change in the average. It is possible that the distribution presentation might elicit higher
11 WTP responses because it may focus respondents' attention on the reduction in the number of
12 relatively bad days (and on the increase in the number of relatively good days) while the
13 associated change in the average may not appear as significant. The implied change in
14 average conditions in the Loehman et al. (1981) San Francisco study turned out to be
15 considerably smaller than that presented in the Brookshire et al. (1982) Los Angeles study,
16 which may have also resulted in a higher value when adjusted to a comparable size change in
17 average visual range because of diminishing marginal utility (i.e., the first incremental
18 improvement is expected to be worth more than the second).
19 The results for the uniform urban haze studies in cities in the eastern United States fall
20 between the respective CVM results for the California cities. The changes in visual range
21 presented in these studies were similar to those presented in the Los Angeles study. In all of
22 the eastern studies respondents were simply asked to consider only the visual effects when
23 answering the WTP questions. This approach is now considered to be inadequate (Irwin
24 etal., 1990; Carson etal., 1990).
25 Irwin et at. (1990) have reported preliminary results for the Denver study (Part H,
26 Table 11-6). The data obtained in this study are being analyzed further and a final report on
27 the project is expected to be available in 1991. Comparison of the results of this study with
28 those from other studies is difficult because the photographs used to illustrate different levels
29 of air quality were not tied to visual range levels. Instead, they were rated on a seven-point
30 air quality scale by the respondents, who were then asked their maximum WTP for a one-step
31 improvement in the scale. This study reports some important methodological findings. One
August 1991
11_53 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
of these is confirmation that simply asking respondents to think only about visibility results in
higher WTP responses than when respondents are asked to give WTP for the change in air
quality and then say what portion of that total is attributable to visibility only. Values for
health and visibility combined exceeded those for visibility alone (when respondents were
simply told to think only about visibility). This may result simply from the effect of budget
constraints on marginal values (the respondent has less to spend on visibility when he also is
buyingihealth); however, the authors express the concern that some, but not all, of the value
for health may be included in the response for visibility. They recommend that respondents
be asked to give total values for changes in urban air quality and then be asked to say what
portion is for visibility.
11.10.4 Conclusions
The current reliance on and continuing methodological uncertainties in the contingent
valuation method for obtaining economic estimates related to changes in visibility means
future methodological research related to CVM may provide important information relevant
to interpreting previous and designing new visibility value studies using CVM. Some
evidence is beginning to emerge as to those factors in CVM applications that appear to have ,
the most effect on visibility value estimates and that, therefore, need to receive more
attention. Among these factors are:
1. The treatment of related air pollution effects such as health;
2. The geographic location of the impact;
3. Whether values are obtained in a single or multiple good context, and are
appropriately adjusted for non-visibility components;
4. Whether a frequency distribution presentation of visibility is used;
5. Various features of the hypothetical context of the WTP questions; and
6. Survey implementation and data handling procedures.
Available estimates of economic values for plumes are directly applicable for only a
thin, dark plume present on most days in the sky at Grand Canyon National Park, and
possibly at a few other national parks in the Southwest. There is little empirical evidence
(from economics studies to date) about how the values would vary with the frequency,
location, or visual characteristics of a plume. Available results do suggest the potential
August 1991
11-54
DRAFT-DO NOT QUOTE OR CITE
-------
1 importance of considering total preservation values, rather than only on-site use values, in
2 any assessment of the economic impact of NOX plumes visible from major recreation areas.
3 The best economic information available for visibility effects associated with NOX is for
4 on-site use values related to changes in visual range in urban areas caused by uniform haze.
5 These values fall roughly between $10 and $100 per year per local household for a 10 percent
6 change in visual range in major urban areas in California and throughout the eastern
7 United States. Reasonable extrapolations of on-site use values (with an order-of-magnitude
8 range of uncertainty) could be made from these studies for estimates of changes in visual
9 range that are attributable to changes in NOX levels in these and other major urban areas,
10 where NO contributes to uniform haze that can be characterized by changes in visual range.
X
11 Extrapolations to less urbanized areas and/or to other visibility changes would require
12 additional assumptions and might introduce additional uncertainty. Because each of the
13 studies completed to date has some important weaknesses and limitations, it would be
14 desirable to. continue to enhance the geographic extent and the technical breadth of issues
15 addressed in these studies to arrive at a broader and more defensible set of estimates.
16 Available results with regard to visual range in urban areas, however, appear to be sufficient
17 to determine the importance of visibility values (on-site use) related to NOx-caused uniform
18 haze in urban areas relative to other potential benefits of NOX controls and to provide order-
19 of-magnitude estimates of such visibility values. To do so would require estimates of the
20 changes in visual range that might be expected as a result of NOX controls.
21 Very little work has been done regarding layered hazes in recreation or residential
22 settings. The work conducted in Denver (Irwin et al., 1990) is the only study in this
23 category. More information is needed about what visual characteristics of such hazes are
24 most important to viewers, as well as on the value people may place on reducing or
25 preventing them. A related question that is relevant for NOX, but that has not been addressed
26 in economic studies to date, is the effect of variations in air pollution color (in plumes or
27 hazes) on visibility values. In order to apply available economic values for changes in
28 layered hazes to changes in NOX levels, a link would need to be made between NOX levels
29 (and changes that might result from control efforts) and visual characteristics of layered hazes
30 that can be tied to the economic values.
August 1991
DRAFT-DO NOT QUOTE OR CITE
-------
1 REFERENCES
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44
45
46
47
48
49
50
51
52
53
Ahlquist, N. C.; Charlson, R. J. (1969) Measurement of the wavelength dependence of atmospheric extinction
due to scatter. Atmos. Environ. 3: 551-564.
Allegrini, I.; De Santis, F. (1989) Measurement of atmospheric pollutants relevant to dry acid deposition Crit
Rev. Anal. Chem. 21: 237-255.
Appel, B. R.; Tokiwa, Y.; Haik, M. (1981) Sampling of nitrates in ambient air. Atmos. Environ. 15: 283-289.
Appel, B. R.; Tokiwa, Y.; Hsu, J.; Kothny, E. L.; Hahn, E.; Wesolowski, J. J. (1983) Visibility reduction as
related to aerosol constituents. Sacramento, CA: State of California, Air Resources Board; report no
CA/DOH/AIHL/SP-29. Available from: NTIS, Springfield, VA; PB84-243617.
Appel, B. R.; Tokiwa, Y.; Hsu, J.; Kothny, E. L.; Hahn, E. (1985) Visibility as related to atmospheric aerosol
constituents. Atmos. Environ. 19: 1525-1534.
Bergstrom, R. W. (1973) Extinktions- und Absorptionskoeffizienten des atmosphaerischen Aerosols als Funktion
der Teilchengroesse [Extinction and absorption coefficients of the atmospheric aerosol as a function of
particle size]. Beitr. Phys. Atmos. 46: 223-234.
Bergstrom, R. W.; Seigneur, C.; Babson, B. L.; Holman, H.-Y.; Wojcik, M. A. (1981) Comparison of the
observed and predicted visual effects caused by power plant plumes. Atmos. Environ. 15: 2135-2150.
Blumenthal, D. L.; Richards, L. W.; Macias, E. S.; Bergstrom, R. W.; Wilson, W. E.; Bhardwaja, P. S. (1981)
Effects of a coal-fired power plant and other sources on southwestern visibility (interim summary of
EPA's project VISTTA). In: White, W. H.; Moore, D. J.; Lodge, J. P., Jr., eds. Plumes and visibility:
measurements and model components, proceedings of the symposium; November 1980; Grand Canyon
National Park, AZ. Atmos. Environ. 15: 1955-1969.
Bravo, A. H.; Saavedra, R. M. I.; Torres, J. R.; Lomas, A. G.; Nava, T. D.; Tirade, S. D. (1988) Particulate
carbon, a significant contributor to the visibility reduction of Mexico City. Geofis. Int. 27: 241-261.
Brookshire, D. S.; Ives, B. C.; Schulze, W. D. (1976) The valuation of aesthetic preference. J. Environ Econ
Manage. 3: 325-346.
Brookshire, D. S.; Thayer, M. A.; Schulze, W. D.; d'Arge, R. C. (1982) Valuing public goods: a comparison
of survey and hedonic approaches. Am. Econ. Rev. 72: 65-83.
Cadle, S. H. (1985) Seasonal variations in nitric acid, nitrate, strong aerosol acidity, and ammonia in an urban
area. Atmos. Environ. 19: 181-188.
Cadle, S. H.; Dasch, J. M.; Mulawa, P. A. (1985) Atmospheric concentrations and the deposition velocity to
snow of nitric acid, sulfur dioxide and various particulate species. Atmos. Environ. 19: 1819-1827.
Carson, R. T.; Mitchell, R. C.; Ruud, P. A. (1990) Valuing air quality improvements: simulating a hedonic
equation in the context of a contingent valuation scenario. In: Mathai, C. V., ed. Visibility and fine
particles: an AWMA/EPA international specialty conference; October 1989; Estes Park, CO. Pittsburgh,
PA: Air and Waste Management Association; pp. 639-646. (AWMA transactions series no. TR-17).
Cass, G. R. (1979) On the relationship between sulfate air quality and visibility with examples in Los Angeles
Atmos. Environ. 13: 1069-1084.
Chandrasekhar, S. (1960) Radiative transfer. Dover, NY:
August 1991
11-56 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18-
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44
45
46
47
48
49
50
51
52
53
54
Charlson, R. J.; Ahlquist, N. C. (1969) Brown haze: NO2 or aerosol? Atmos. Environ. 3: 653-656.
Charlson, R. J.; Covert, D. S.; Tokiwa, Y.; Mueller, P. K. (1972) Multiwavelength nephelometer measurements
in Los Angeles smog aerosol: III. comparison to light extinction by NO2. In: Hidy, G. M., ed. Aerosols
and atmospheric chemistry: the Kendall award symposium at the proceedings of the American Chemical
Society; March-April 1971; Los Angeles, CA. New York, NY: Academic Press, Inc.; pp. 333-338.
Chestnut, L. G.; Rowe, R. D. (1989) Economic valuation of changes in visibility: a state of the science
assessment for NAPAP. In: National acid precipitation assessment program: methods for valuing acidic
deposition and air pollution effects. Washington, DC: National Acid Precipitation Assessment Program;
SOS/T report no. 27.
Chestnut, L. G.; Rowe, R. D. (1990) Preservation values for visibility in the national parks. Washington, DC:
U. S. Environmental Protection Agency.
Colbeck, I.; Harrison, R. M. (1984) Ozone-secondary aerosol-visibility relationships in north-west England. Sci.
Total Environ. 34: 87-100.
Cornsweet, T. N. (1970) Factors other than wavelength that influence hue. In: Visual perception. New York,
NY: Academic Press, Inc.; pp. 236-240.
Cummings, R. G.; Brookshire, D. S.; Schulze, W. D., eds. (1986) Valuing environmental goods: an assessment
of the contingent valuation method. Totowa, NJ: Rowman and Allanheld.
Daniel, T. C.; Hill, A. C. (1987) Measuring visibility values: comparison of perceptual assessment methods. In:
' Bhardwaja, P. S., ed. Visibility protection: research and policy aspects. Pittsburgh, PA: Air Pollution
Control Association.
Diederen, H. S. M. A.; Guicherit, R.; Hollander, J. C. T. (1985) Visibility reduction by air pollution in The
Netherlands. Atmos. Environ. 19: 377-383.
Dixon, J. K. (1940) The absorption coefficient of nitrogen dioxide in the visible spectrum. J. Chem. Phys. 8:
157-160.
Drivas, P. J.; Machiraju, S.; Heinold, D. W. (1980) ERT visibility model: version 3, technical description and
user's guide. Concord, MA: Environmental Research & Technology.
Dunwoody, C. L. (1986) Rapid nitrate loss from PM10 filters. J. Air Pollut. Control Assoc. 36: 817-818.
Dzubay, T. G.; Clubb, K. W. (1981) Comparison of telephotometer measurements of extinction coefficients with
scattering and absorption coefficients. In: White, W. H.; Moore, D. J.; Lodge, J. P., Jr., eds. Plumes
and visibility: measurements and model components, proceedings of the symposium; November 1980;
Grand Canyon National Park, AZ. Atmos. Environ. 15: 2617-2624.
Dzubay, T. G.; Stevens, R. K.; Lewis, C. W.; Hern, D. H.; Courtney, W. J.; Tesch, J. W.; Mason, M. A.
(1982) Visibility and aerosol composition in Houston, Texas. Environ. Sci. Technol. 16: 514-525.
Eltgroth, M. W. (1982) A comparison of measured absolute plume intensities versus predictions by the
PHOENIX and PLUVUE models. Austin, TX: Radian Corporation; Radian document no.
141-485-2-047.
Ferman, M. A.; Wolff, G. T.; Kelly, N, A. (1981) The nature and sources of haze in the Shenandoah
Valley/Blue Ridge Mountains area. J. Air Pollut. Control Assoc. 31: 1074-1082.
August 1991
11-57
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44
45
46
47
48
49
50
Finlan, J. M. (1981) Scattering of light and the color of atmospheric hazes. In: White, W. H.; Moore, D. J.;
Lodge, J. P., Jr., eds. Plumes and visibility: measurements and model components, proceedings of the
symposium; November 1980; Grand Canyon National Park, AZ. Atmos. Environ. 15: 2599-2616.
Fischhoff, B.; Furby, L. (1988) Measuring values: a conceptual framework for interpreting transactions with
special reference to contingent valuation of visibility. J. Risk Uncertainty 1: 147-184.
Freeman, A. M., HI. (1979) The benefits of environmental improvement: theory and practice. Baltimore, MD:
The Johns Hopkins University Press.
Graves, P.; Murdoch, J. C.; Thayer, M. A.; Waldman, D. (1988) The robustness of hedonic price estimation:
urban air quality. Land Econ. 64: 220-233.
Gray, H. A.; Cass, G. R.; Huntzicker, J. J.; Heyerdahl, E. K.; Rau, J. A. (1984) Elemental and organic carbon
particle concentrations: a long-term perspective. Sci. Total Environ. 36: 17-25.
Groblicki, P. J.; Wolff, G. T.; Countess, R. J. (1981) Visibility-reducing species in the Denver "brown cloud"—
L relationships between extinction and chemical composition. In: White, W. H.; Moore, D. J.; Lodge,
J. P., Jr., eds. Plumes and visibility: measurements and model components, proceedings of the
symposium; November 1980; Grand Canyon National Park, AZ. Atmos. Environ. 15: 2473-2484.
Haas, P. J.; Fabrick, A. J. (1981) The effects of NO2-aerosol interaction on indices of perceived visibility
impairment. Atmos. Environ. 15: 2171-2177.
Hasan, H.; Dzubay, T. G. (1983) Apportioning light extinction coefficients to chemical species in atmospheric
aerosol. Atmos. Environ. 17: 1573-1581.
Hegg, D. A.; Hobbs, P. V. (1983) Particles and trace gases in the plume from a modern coal-fired power plant
in the western United States and their effects on light extinction. Atmos. Environ. 17: 357-368.
Hodkinson, J. R. (1966) Calculations of colour and visibility in urban atmospheres polluted by gaseous NO2. Int
J. Air Water Pollut. 10: 137-144.
Horvath, H. (1971) On the brown colour of atmospheric haze. Atmos. Environ. 5: 333-344.
Husar, R. B.; White, W. H. (1976) On the color of the Los Angeles smog. Atmos. Environ. 10: 199-204.
Invin, J.; Schenk, D.; McClelland, G. H.; Schulze, W. D.; Stewart, T.; Thayer, M. (1990) Urban visibility:
some experiments on the contingent valuation method. In: Mathai, C. V., ed. Visibility and fine
particles: an AWMA/EPA international specialty conference; October 1989; Estes Park, CO. Pittsburgh,
PA: Air and Waste Management Association; pp. 647-658. (AWMA transactions series no. TR-17).
John, W. (1986) A new method for nitric acid and nitrate aerosol measurement using the dichotomous sampler.
Sacramento, CA: California Air Resources Board; GARB document n'o. CA/DOH/AIHL/R-304.
John, W.; Wall, S. M.; Ondo, J. L. (1985) Dry acid deposition on materials and vegetation: concentrations in
ambient air. Berkeley, CA: California Department of Health Services, Air and Industrial Hygiene
Laboratory; report no. CA/DOH/AIHL/SP-34.
August 1991
11-58
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44
45
46
47
48
49
50
51
52
53
54
Johnson R. W. (1981) Daytime visibility and nephelometer measurements related to its determination. In: White,
W H • Moore D J.; Lodge, J. P., Jr., eds. Plumes and visibility: measurements and model
components, proceedings of the symposium; November 1980; Grand Canyon National Park, AZ. Atmos.
Environ. 15: 1835-1845.
Johnson, C. D.; Latimer, D. A.; Bergstrom, R. W.; Hogo, H. (1980) User's manual for the plume visibility
model (PLUVUE). Research Triangle Park, NC: U. S. Environmental Protection Agency, Office of Air
Quality Planning and Standards; EPA report no. EPA-450/4-80-032. Available from: NTIS, Springfield,
VA; PB81-163297.
Kerker, M. (1969) The scattering of light and other electromagnetic radiation. New York, NY: Academic Press,
' Inc. (Loebl, E. M., ed. Physical chemistry: a series of monographs, no. 16).
Koschmieder, H. (1924) Theorie der horizontalen Sichtweite II: Kontrast und Sichtweite [Theory of horizontal
visibility]. Beitr. Phys. Atmos. 12: 171-181.
Latimer, D. A. (1979) Power plant,impacts on air quality and visibility: siting and emission control implications.
Washington, DC: U. S. Environmental Protection Agency, Office of Planning and Evaluation; EPA
publication no. EPA-230/11-79-001. Available from: NTIS, Springfield, VA; PB80-123516.
Latimer, D. A. (1980) Power plant impacts on visibility in the West: siting and emissions control implications.
J. Air Pollut. Control Assoc. 30: 142-146.
Latimer, D. A. (1988a) Assessing the effects of changes in aerosol concentrations on seasonal and annual visual
range. San Rafael, .CA: Gaia Associates.
Latimer D A (1988b) Plume perceptibility thresholds. Presented at: 81st annual meeting of the Air Pollution
Control Association; June; Dallas, TX. Pittsburgh, PA: Air Pollution Control Association; paper no.
88-56.6.
Latimer, D. A. (1988c) Application of disperion models for layered haze source apportionment. Presented at:
81st annual meeting of the Air Pollution Control Association; June; Dallas, TX. Pittsburgh, PA: Air
Pollution Control Association; paper no. 88-52.6.
Latimer, D. A.; Ireson, R. G. (1980) Workbook for estimating visibility impairment. Research Triangle Park,
NC: U. S. Environmental Protection Agency, Office of Air Quality Planning and Standards; EPA
publication no. EPA-450/4-80-031. Available from: NTIS, Springfield, VA; PB81-157885.
Latimer D. A.; Ireson, R. G. (1988) Workbook for plume visual impact screening and analysis. Research
/Triangle Park, NC: U. S. Environmental Protection Agency, Office of Air Quality Planning and
Standards; EPA report no. EPA-450/4-88-015. •>
Latimer, D. A.; Samuelsen, G. S. (1975) Visual impact of plumes from power plants. Irvine, CA: University of
California, School of Engineering, Air Quality Laboratory; publication no. UCI-ARTR-75-3.
Latimer, D. A.; Samuelsen, G. S. (1978) Visual impact of plumes from power plants: a theoretical model.
Atmos. Environ. 12: 1455-1465.
Latimer, D. A.; Bergstrom, R. W.; Hayes, S. R.; Liu, M.-K.; Seinfeld, J. H.; Whiten, G. Z.; Wojcik, M. A.;
Hillyer M. J. (1978) The development of mathematical models for the prediction of anthropogenic
visibility impairment: volumes I-III. Research Triangle Park, NC: U. S. Environmental Protection
Agency, Office of Air Quality Planning and Standards; EPA report no. EPA-450/3-78-1 lOa-c. Available
from: NTIS, Springfield, VA; PB-293118-SET.
August 1991
11_59 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44
45
46
47
48
49
50
51
52
53
Latimer, D. A.; Hogo, H.; Daniel, T. C. (1981) The effects of atmospheric optical conditions on perceived
scenic beauty. Atmos. Environ. 15: 1865-1874.
Latimer, D. A.; Hogo, H.; Chinkin, L. R.; Dudik, M. C.; Ireson, R. G.; Irpan, P.; Jacobson, H.; Killus, J. P.;
Lundberg, G. W.; Saxena, P.; Whitten, G. Z.; Yocke, M. A. (1985a) Modeling regional haze in the
Southwest: a preliminary assessment of source contributions, volume 1 [revised draft report]. San Rafael,
CA: Systems Applications, Inc.; chps. 4 and 5; SAI report no. SYSAPP/85-038.
Latimer, D. A.; Chinkin, L. R.; Dudik, M. C.; Hogo, H.; Ireson, R. G. (1985b) Uncertainties associated with
modeling regional haze in the Southwest. Washington, DC: American Petroleum Institute, Health and
Environmental Sciences Department, API publication no. 4403.
Latimer, D. A.; Gery, M. W.; Hogo, H. (1986) A theoretical evaluation of the role of nighttime nitrate
formation in the formation of layered haze. San Rafael, CA: Systems Applications, Inc.; SAI document
no. SYSAPP-86-167.
Lewis, C. W.; Baumgardner, R. E.; Stevens, R. K.; Russwurm, G. M. (1986) Receptor modeling study of
Denver winter haze. Environ. Sci. Technol. 20: 1126-1136.
Loehman, E.; Boldt, D.; Chaikin, K. (1981) Measuring the benefits of air quality improvements in the San
Francisco Bay area. Washington, DC: U. S. Environmental Protection Agency.
MacFarland, K. K.; Malm, W.; Molenar, J. (1983) An examination of methodologies and social indicators for
assessing the value of visibility. la: Rowe, R. D.; Chestnut, L. G., eds. Managing air quality and scenic
resources at national parks and wilderness areas. Boulder, CO: Westview Press.
Macias, E. S.; Zwicker, J. O.; Ouimette, J. R.; Hering, S. V.; Friedlander, S. K.; Cahill, T. A.; Kuhlmey,
G. A.; Richards, L. W. (1981) Regional haze case studies in the southwestern U.S.— I. aerosol chemical
composition. In: White, W. H.; Moore, D. J.; Lodge, J. P., Jr., eds. Plumes and visibility:
measurements and model components, proceedings of the symposium; November 1980; Grand Canyon
National Park, AZ. Atmos. Environ. 15: 1971-1986.
Malm, W. C.; Gebhard, K. A. (1988) Optical characteristics of aerosols at three national parks. Presented at:
81st annual meeting of the Air Pollution Control Association; June; Dallas, TX. Pittsburgh, PA: Air
Pollution Control Association; paper no. 88-52.2.
Malm, W. C.; Iyer, H. K. (1988) Examination of the relationship between Navajo Generating Station emissions
and aerosol concentrations at Page, Arizona. Presented at: 81st annual meeting of the Air Pollution
Control Association; June; Dallas, TX. Pittsburgh, PA: Air Pollution Control Association; paper no
88-52.4.
Malm, W.; Gebhard, K.; Cahill, T.; Eldred, R.; Pielke, R.; Watson, J.; Latimer, D. (1989) Winter haze
intensive experiment: draft final report. Fort Collins, CO: National Park Service.
Mamane, Y.; Dzubay, T. G. (1986) Characteristics of individual particles at a rural site in the eastern United
States. J. Air Pollut. Control Assoc. 36: 906-911.
Mamane, Y.; Mehler, M. (1987) On the nature of nitrate particles in a coastal urban area. Atmos. Environ 21-
1989-1994. ' '
Mathai, C. V.; Tombach, I. H. (1987) A critical assessment of atmospheric visibility and aerosol measurements
in the eastern United States. J. Air Pollut. Control Assoc. 37: 700-707.
August 1991
11-60 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24 .
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44
45
46
47
48
49
50
51
52
53
54
Melo, O. T.; Stevens, R. D. S. (1981) The occurrence and nature of brown plumes in Ontario. Atmos. Environ.
15: 2521-2529.
Middleton, W. E. K. (1952) Vision through the atmosphere. Toronto, ON, Canada: University of Toronto Press.
Middleton, P.; Stewart, T. R.; Dennis, R. L. (1983) Modeling human judgments of'urban visual air quality.
Atmos. Environ. 17: 1015-1021.
Milford, J. B.; Davidson, C. I. (1987) The sizes of participate sulfate and nitrate in the atmosphere—a review.
J. Air Pollut. Control Assoc. 37: 125-134.
Mitchell, R. C.; Carson, R. T. (1989) Using surveys to value public goods: the contingent valuation method.
Washington, DC: Resources for the Future.
Mulawa, P. A.; Cadle, S. H. (1985) A comparison of nitric acid and particulate nitrate measurements by the
penetration and demider difference methods. Atmos. Environ. 19: 1317-1324.
Orel, A. E.; Seinfeld, J. H. (1977) Nitrate formation in atmospheric aerosols. Environ. Sci Technol'. 11:
1000-1007.
Pierson, W. R.; Brachaczek, W. W.; Gorse, R. A., Jr.; Japar, S. M.; Norbeck, J. M.; Keeler, G. J. (1987)
Acid rain and atmospheric chemistry at Allegheny Mountain. Environ. Sci. Technol. 21: 679-691.
Pratsinis, S.; Ellis, E. C.; Novakov, T.; Friedlander, S. K. (1984) The carbon containing component of the Los
Angeles aerosol: source apportionment and contributions to the visibility budget. J. Air Pollut. Control
Assoc. 34: 643-650.
Rae, D; A. (1983) The value to visitors of improving visibility at Mesa Verde and Great Smoky National Parks.
In: Rowe, R. D.; Chestnut, L. G., eds. Managing air quality and scenic resources at national parks and
wilderness areas. Boulder, CO: Westview Press.
Rae, D. A. (1984) Benefits of visual air quality in Cincinnati - results of a contingent ranking survey, final
report. Palo Alto, CA: Electric Power Research Institute; report no. RP-1742.
Richards, L. W. (1983) Comments on the oxidation of NO2 to nitrate—day and night. Atmos. Environ. 17:
397-402.
Richards, L. W.; Anderson, J. A.; Blumenthal, D. L.; Brandt, A. A.; McDonald, J. A.; Waters, N.; Macias,
E. S.;'Bhardwaja, P. S. (1981) The chemistry, aerosol physics, and optical properties of a western
coal-fired power plant plume. Atmos. Environ. 15: 2111-2134.
Robinson, E. (1968) Effect on the physical properties of the atmosphere. In: Stern, A. C., ed. Air pollution:
v. 1, air pollution and its effects. 2nd ed. New York, NY: Academic Press, Inc.; pp. 349-400.
Ross, D. M.; Malm, W. C.; Loomis, R. J. (1987) An examination of the relative importance of park attributes
at several national parks. In: Bhardwaja, P. S., ed. Visibility protection: research and policy aspects.
Pittsburgh, PA: Air Pollution Control Association.
Rowe, R. D.; Chestnut, L. G. (1982) The value of visibility: economic theory and applications for air pollution
control. Cambridge, MA: Abt Books.
Russell, A. G.; Cass, G. R. (1986) Verification of a mathematical model for aerosol nitrate and nitric acid
formation and its use for control measure evaluation. Atmos. Environ. 20: 2011-2025.
August 1991
11-61
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44
45
46
47
48
49
50
51
52
53
54
Russell, A. G.; McCue, K. F.; Cass,.G. R. (1988) Mathematical modeling of the formation of
nitrogen-containing air pollutants. 1, Evaluation of an Eulerian photochemical model. Environ. Sci.
Technol. 22: 263-271.
Ruud, P. A. (1987) Contingent ranking surveys: their application and design in estimating the value of visibility.
In: Bhardwaja, P. S., ed. Visibility protection: research and policy aspects. Pittsburgh, PA: Air Pollution
Control Association.
SAROAD Data, Storage and Retrieval of Aerometric Data [data base], (n.d.) [Data files for nitrogen dioxide,
1980 through 1987; and for coarse and fine (inhalable) particulate (IP), 1979 through 1983. Data bases
, were converted in 1987 to AIRS, Aerometric Information Retrieval System]. Research Triangle Park,
NC: U. S. Environmental Protection Agency, Office of Air Quality Planning and Standards; Disk-
ASCH. , • ,- ..
Saxena, P.; Hudischewskyj, A. B.; Seigneur, C.; Seinfeld, J. H. (1986) A comparative study of equilibrium
approaches to the chemical characterization of secondary aerosols. Atmos. Environ. 20: 1471-1483.
Schulze, W. D.; Brookshire, D. S.; Walther, E.; MacFarland, K. K.; Thayer, M. A.; Whitworth, R. L.;
Ben-David, S.; Malm, W.; Molenar, J. (1983) The benefits of preserving visibility in the national
parklands of the southwest. Nat. Resour. J. 23: 149-173.
Seigneur, C.; Johnson, C. D.; Latimer, D. A,; Bergstrom, R. W.; Hogo, H. (1984) User's manual for the plume
visibility model (PLUVUE H). Research Triangle Park, NC: U. S. Environmental Protection Agency,
Environmental Sciences Research Laboratory; EPA report no. EPA-600/8-84-005. Available from- NTIS
Springfield, VA; PB84-158302.
Shah, J. J.; Watson, J. G., Jr.; Cooper, J. A.; Huntzicker, J. J. (1984) Aerosol chemical composition and light
scattering in Portland, Oregon: the role of carbon. Atmos. Environ. 18: 235-240.
Sloane, C. S. (1987) The contribution of NO2 and soot to the discoloration of urban skies. In: Bhardwaja, P. S.,
ed. Visibility protection: research and policy aspects, an APCA specialty conference; September 1986;
Grand Teton National Park, WY; Pittsburgh, PA: Air Pollution Control Association; pp. 434-452.
(APCA transactions no, TR-10). .,
Sloane, C. S.; White, W. H. (1986) Visibility: an evolving issue. Environ. Sci. Technol. 20: 760-766.
Solomon, P. A.; Movers, J. L. (1984) Use of a high volume dichotomous virtual impactor to estimate light
extinction due to carbon and related species in the Phoenix haze. Sci. Total Environ. 36: 169-175.
Solomon, P. A.; Moyers, J. L. (1986) A chemical characterization of wintertime haze in Phoenix, Arizona.
Atmos. Environ. 20: 207-213.
Stelson, A. W. ; Seinfeld, J. H. (1982) Relative humidity and temperature dependence of the ammonium nitrate
dissociation constant. Atmos. Environ. 16: 983-992.
Stelson, A. W.; Friedlander, S. K.; Seinfeld, J. H. (1979) A note on the equilibrium relationship between
ammonia and nitric acid and particulate ammonium nitrate. Atmos. Environ. 13: 369-371.
Stevens, R. K. (1987) [Personal communication to J. C. Trijonis]. Research Triangle Park, NC: U. S.
Environmental Protection Agency, Atmospheric Research and Exposure Assessment Laboratory [as cited •
in Trijonis (1987)]. .
Stevens, R. K.; Dzubay, T. G.; Shaw, R. W., Jr.; McClenny, W. A.; Lewis, C. W.; Wilson, W. E. (1980)
Characterization of the aerosol in the Great Smoky Mountains. Environ. Sci. Technol. 14: 1491-1498.
August 1991
11-62 DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22;
23
24 .
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44
45
46
47
48
49
50
51
52
53
54
Stevens, R. K.; King, F.; Bell, J.; Whitfield, J. (1988) Measurement of the chemical species that contribute to
urban haze. Presented at: 81st annual meeting .of the Air Pollution Control Association; June; Dallas,
TX. Pittsburgh, PA: Air Pollution Control Association; paper no. 88-57.3.
Tang, I. N.; Wong, W. T.; Munkelwitz, H. R. (1981) The relative importance of atmospheric sulfates and
nitrates in visibility reduction. Atmos. Environ. 15: 2463-2471.
Tolley, G. A.; Randall, A.; Blomquist, G.; Fabian, R.; Fishelson, G.; Frankel, A.; Hoehn, J.; Krumm, R.;
Mensah, E.; Smith, T. (1986) Establishing and valuing the effects of improved visibility in eastern
United States. Washington, DC: U. S. Environmental Protection Agency.
Tombach, I. H.; Allard, D. W.; Drake, R; L.; Lewis, R. C. (1987) Western regional air quality studies,
visibility and air quality measurements: 1981 and 1982, interim report. Palo Alto, CA: Electric Power
Research Institute; report no. EA-4903, research project 1630-11.
Trijonis, J. (1982a) Existing and natural background levels of visibility and fine particles in the rural East.
Atmos. Environ. 16: 2431-2445.
Trijonis, J. (1982b) Visibility in California. J. Air Pollut. Control Assoc. 32: 165-169.
Trijonis, J. (1987) National relationship between visibility and NOX emissions. Bloomington, MN: Santa Fe
Research Corporation; EPA contract no, 6W-3191-NASX.
Trijonis, J.; Davis, M. (1981) Development and application of methods for estimating inhalable and fine particle
concentrations from routine hi-vol data. Sacramento, CA: State of California, Air Resources Board;
GARB report no. ARB/R-81/171. Available from: NTIS, Springfield, VA; PB83-117572.
Trijonis, J.; Yuan, K. (1978a) Visibility in the Northeast: long-term visibility trends and visibility/pollutant
relationships. Research Triangle Park, NC: U. S. Environmental Protection Agency, Environmental
" , Sciences Research Laboratory; EPA report no. EPA-600/3-78-075. Available from: NTIS, Springfield,
.VA; PB-286921. •
Trijonis, J.; Yuan, K. (1978b) Visibility in the southwest: an exploration of the historical data base. Research
Triangle Park, NC: U. S. Environmental Protection Agency, Environmental Sciences Research
Laboratory; EPA report no. EPA-600/3-78-039. Available from: NTIS, Springfield, VA; PB-282942.
Trijonis, J.; Cass, G.; McRae, G.; Horie, Y.; Lim, W.-Y.; Chang, N.;-Cahill, T. (1982) Analysis of
visibility/aerosol relationships and visibility modeling/monitoring alternatives for California. Sacramento,
CA: State of California, Air Resources Board; report no. ARB-R-83-197. Available from: NTIS;
Springfield, VA; PB83-245563.
Trijonis, J.; Thayer, M.; Murdoch, J.; Hageman, R. (1984) Air quality benefits analysis for Los Angeles and
San Francisco based on housing values and visibility. Sacramento, CA: California Air Resources Board.
Trijonis, J. C.; Pitchford, M.; McGown, M. (1987) Preliminary extinction budget results from the RESOLVE
program. In: Bhardwaja, P. S., ed. Visibility protection: research and policy aspects, an APCA specialty
conference; September 1986; Grand Teton National Park, WY. Pittsburgh, PA: Air Pollution Control
Association; pp. 872-883. (APCA transactions no. TR-10).
Trijonis, J.; McGown, M.; Pitchford, M.; Blumenthal, D.; Roberts, P.; White, W.; Macias, E.; Weiss, R.;
Waggoner, A.; Watson, J.; Chow, J.; Flocchini, R. (1988) Visibility conditions and causes of visibility
degradation in the Mojave Desert of California: executive summary, RESOLVE project final report.
China Lake, CA: Department of the Navy, Naval Weapons Center; document no. NWC TP 6869.
Available from: NTIS, Springfield, VA; AD-A206 322.
August 1991
11-63
DRAFT-DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44
45
46
47
48
49
50
51
52
53
54
U. S. Environmental Protection Agency. (1979) Protecting visibility; an EPA report to Congress. Research
Triangle Park, NC: Office of Air Quality Planning and Standards; EPA report no. EPA-450/5-79-008.
Available from: NTIS, Springfield, VA; PB80-220320.
U. S. Environmental Protection Agency. (1985) Developing long-term strategies for regional haze: findings and
recommendations of the visibility task force. Research Triangle Park, NC.
Vanderpol, A. H.; Humbert, M. E. (1981) Coloration of power plant plumes—NO2 or aerosols? Atmos, Environ.
15: 2105-2110.
Vossler, T. L.; Lewis, C. W.; Stevens, R. K.; Dzubay, T. G.; Gordon, G. E.; Tuncel, S. G.; Russwurm,
G. M.; Keeler, G. J. (1989) Composition and origin of summertime air pollutants at Deep Creek Lake,
Maryland. Atmos. Environ. 23: 1535-1547.
Waggoner, A. P.; Weiss, R. E.; Ahlquist, N. C. (1983) The color of Denver haze. Atmos. Environ. 17:
2081-2086.
Wall, S. M.; John, W.; Ondo, J. L. (1988) Measurement of aerosol size distributions for nitrate and major ionic
species. Atmos. Environ. 22: 1649-1656.
White, W. H. (1981) The role of participate sulfates and nitrates in reducing visibility. Presented at: 74th annual
meeting of the Air Pollution Control Association; June; Philadelphia, PA. Pittsburgh, PA: Air Pollution
Control Association; paper no. 81-54.1.
White, W. H. (1982) Effects of nitrogen oxides on visibility. In: Air quality criteria for oxides of nitrogen.
Research Triangle Park, NC: U. S. Environmental Protection Agency, Office of Health and
Environmental Assessment, Environmental Criteria and Assessment Office; EPA report no.
EPA-600/8-82-026. Available from: NTIS, Springfield, VA; PB83-131011.
White, W. H.; Macias, E. S. (1987) Particulate nitrate measurements in rural areas of the western United States.
Atmos. Environ. 21: 2563-2571.
White, W. H.; Patterson, D. E. (1981) On the relative contributions of NO2 and particles to the color of smoke
plumes. Atmos. Environ. 15: 2097-2104.
White, W. H.; Roberts, P. T. (1977) On the nature and origins of visibility-reducing aerosols in the Los Angeles
air basin. Atmos. Environ. 11: 803-812.
White, W. H.; Seigneur, C.; Heinold, D. W.; Eltgroth, M. W.; Richards, L. W.; Roberts, P. T.; Bhardwaja,
P. S.; Conner, W. D.; Wilson, W. E., Jr. (1985) Predicting the visibility of chimney plumes: an
intercomparison of four models with observations at a well-controlled power plant. Atmos. Environ. 19:
515-528.
White, W. H.; Seigneur, C.; Heinold, D. W.; Richards, L. W.; Wilson, W. E.; Roberts, P. T. (1986) Radiative
transfer budgets for scattering and absorbing plumes: measurements and model predictions. Atmos.
Environ. 20: 2243-2257.
Williams, M. D.; Treiman, E.; Wecksung, M. (1980) Plume blight visibility modeling with a simulated
photograph technique. J. Air Pollut. Control Assoc. 30: 131-134.
Williams, M.; Chan, L. Y.; Lewis, R. (1981) Validation and sensitivity of a simulated-photograph technique for
visibility modeling. Atmos. Environ. 15: 2151-2170.
Wolff, G. T. (1984) On the nature of nitrate in coarse continental aerosols. Atmps. Environ. 18: 977-981.
August 1991
11-64 DRAFT-DO NOT QUOTE OR CITE
-------
1
2.
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
Wolff, G. T.; Korsog, P. E. (1989) Atmospheric concentrations and regional source apportionments of sulfate,
1 nitrate and sulfur dioxide in the Berkshire Mountains in western Massachusetts. Atmos. Environ. 23:
55-65. '
Wolff, G. T.; Countess, R. J.; Groblicki, P. J.; Ferman, M. A.; Cadle, S. H.; Muhlbaier, J. L. (1981)
Visibility-reducing species in the Denver "brown cloud"— II. sources and temporal patterns. Atmos.
Environ. 15: 2485-2502.
Wolff, G. T.; Ferman, M. A.; Kelly, N. A.; Stroup, D. P.; Ruthkosky, M. S. (1982) The relationships between
the chemical composition of fine particles and visibility in the Detroit metropolitan area. J. Air Pollut.
Control Assoc. 32: 1216-1220,
Yoshizumi, K. (1986) Regional size distributions of sulfate and nitrate in the Tokyo metropolitan area in
summer. Atmos. Environ. 20: 763-766.
Zak, B. D.; Einfeld, W.; Church, H. W.; Gay, G. T.; Jensen, A. Lf; Trijonis, J.; Ivey, M. D.; Homann, P. S.;
Tipton, C. (1984) The Albuquerque winter visibility study. Volume 1: overview and data analysis.
Albuquerque, NM: U. S. Department of Energy, Sandia National Laboratories; Sandia document no.
SAN84-0173/1. Available from: NTIS, Springfield, VA; DE84014356.
August 1991
H-65 ' DRAFT-DO NOT QUOTE OR CITE
-------
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
12. EFFECTS OF NITROGEN OXIDES
ON MATERIALS
12.1 INTRODUCTION
Materials exposed to the atmosphere in both indoor and outdoor environments may
suffer undesirable physical and chemical changes. Although many of these changes occur
whether or not pollutants are present, the rate at which these changes occur can be influenced
by pollutant concentrations. Nitrogen oxides (NOX), including nitric oxide (NO), nitrogen
dioxide (NO2), and nitric acid (HNO3), are known to affect the fading of dyes; the strength
of fabrics, plastics, and rubber products; the corrosion of metals; and the use-life of
electronic components, paints, and masonry. While the materials damage potential of sulfur
oxides (SOX) has been extensively studied, much less research has been reported for NOX.
Graedel and McGill (1986) have pointed out, however, that sulfur dioxide (SO2)
concentrations are generally decreasing across the country and NOX levels are increasing.
The relative proportion of materials damage attributable to NOX can therefore be expected to
increase. This section discusses the impact of NOX on a number of categories of materials.
Emphasis is placed on those experiments and materials in which degradation was observed.
To understand the results of materials exposure to NOX, it is important to appreciate the
influence of several factors on the materials damage process:
1. The environment in which materials are exposed;
2. The mechanisms that cause damage in different exposures;
3. The wet and dry deposition processes that influence damage rates; and
4. The chemical interactions of NOX species with materials and with other
components of the environment; for example, other airborne pollutants and
moisture.
It is also necessary to understand the experimental techniques used to study damage processes
and the limitations of these study techniques, as well as the results of the studies. Finally, if
estimates of the costs of materials damage are desired, a qualitative understanding of the
August 1991
12-1 DRAFT - DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
economic estimation procedures is needed. A useful survey of the topic of air pollution
damage to materials is contained in Joerg et al. (1987).
12.1.1 Environmental Exposures to Materials
The materials affected by NOX occur in both indoor and outdoor environments.
Outdoor materials will be exposed to NOX concentrations such as those discussed in Chapter
7 plus stresses caused by a wide range of temperatures and humidities, sunlight, and
precipitation. Identical materials exposed in nearby locations may be damaged at very
different rates depending on their microenvironments (e.g., building stone sheltered by an
overhang will be damaged at a different rate than stone openly exposed on the face of the
same structure). Most materials exposed for extended periods to the outdoor environment are
selected or designed to withstand these exposures and, therefore, they degrade at a slow rate.
Materials that may be subject to NOX damage and that are widely used outdoors include
paints; cement and concrete; stone; architectural and statuary metals; plastics; and elastomers.
Indoor concentrations of NOX are discussed in Chapter 7. Although indoor
environments are free of many of the extreme environmental stresses present outdoors, NOX
concentrations may be significantly higher in some indoor environments (e.g., where
unvented gas appliances are in use) and the materials exposed indoors may be more sensitive.
Virtually all the materials found outside are also found indoors to some extent; however,
additional materials such as paper, fine textiles, and electronic components are more common
in indoor than outdoor environments. In addition, paint formulations intended for indoor
applications are different from those formulations intended for outdoor use.
12.1.2 Mechanisms of Materials Damage
Damage to exposed materials results from attack through both physical and chemical
processes, and damage is induced both by pollution and other agents. Physical processes
include erosion by windborne particles, differential heating, and frost attack. Chemical
processes include corrosion, biological attack (e.g., mildew), direct attack by acid mists, and
gaseous and particle deposition and subsequent reactions (Tombach, 1982; Yocom and Baer,
1983). It is difficult to distinguish a single causative agent for observed damage to exposed
materials because many agents, together with a number of environmental stresses, act on a
August 1991
12-2 DRAFT - DO NOT QUOTE OR CITE
-------
1 surface throughout its life. Even extensively studied systems (such as the effect of SO2
2 pollution on metals) are not thoroughly understood, and there is extensive work still needed
3 to understand the interaction of NOX with the variety of materials in use today.
4 .
5 12.1.3 Deposition Processes
6 For them to cause damage to a material, atmospheric NOX must come in contact with ,
7 the material. They are deposited on material surfaces through both wet and dry deposition
8 processes (Tombach, 1982). Dry deposition processes for gaseous NOX include Brownian or
9 molecular diffusion to the surface, Stefan flow toward surfaces where moisture is condensingj
10 thermophoresis toward cold surfaces, and diffusiophoresis toward evaporating surfaces. In
11 addition, particles containing NOX can be transported to a material surface through
12 gravitational settling or inertia! impaction of the particles on the surface. Wet deposition
13 (e.g., acid rain) processes include the scavenging of gaseous NOX or particles containing
14 absorbed NOX into precipitation or fog droplets that impact the surface. The rate at which
15 deposition processes transport NOX to the surface is dependent on the NOX concentrations in
16 the environment, the chemistry and geometry of the surface, the concentrations of other
17 atmospheric constituents, and the turbulent transfer properties of the air (Lipfert, 1989).
18 The transfer of pollutants from the atmosphere to a surface is often visualized in terms
19 of the "multiple resistance analogy" (Sherwood et al., 1990). In this analogy, the rate of
20 mass transfer of pollutants is modeled as being determined by a series of resistances to the
21 mass transfer. The total resistance, RT, is made up of the sum of "free air" turbulent transfer
22 resistance, Ra; the near-surface, quasi-laminar boundary layer resistance, Rb; and the surface
23 uptake resistance, Rc. Thus,
24 RT = Ra + Rb + Rc. (12-1)
25
26 The aerodynamic resistance, Ra, is dominated by atmospheric turbulence. The
27 boundary layer resistance, Rb, depends on the aerodynamics of flow immediately adjacent to
28 the surface and the molecular diffusivity of the pollutant. The surface resistance, Rc, depends
29 on the physical and chemical interactions of the surface and the pollutant. Depending on the
30 aerodynamic conditions, and the physical and chemical state of the surface, any of these
31 terms can be the rate-limiting step for the transfer.
August 1991
12-3 DRAFT - DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
The inverse of the total resistance is the deposition velocity, Vd, which has units of
cm/s. The deposition velocity is the ratio of flux of mass to the surface (g/cm2 s) to the free
air concentration of the pollutant (g/cm3).
In a laboratory study, Edney et al. (1986) measured the deposition of NO2 and various
other compounds to both wet and dry galvanized steel. A large "smog chamber" (an
environmental chamber designed to simulate photochemical processes) was used for the study
and NO2, propylene (C3H3), and SO2 were introduced in various combinations to study
deposition processes. Galvanized steel was exposed both dry and wet with artificial dew
cycles caused by cooling the samples. An experiment with a dry surface and NO2 alone
yielded a velocity for the deposition of NO2 to galvanized steel of 0.05 cm/s. A similar test
with SO2 yielded an SO2-to-galvanized steel deposition velocity of 0.8 cm/s, or deposition
about sixteen times greater for S02 than for NO2. Clearly, dry deposition of NO2 on
galvanized steel is significantly slower than the dry deposition of SO2. These researchers
suggest that, for the purposes of developing a damage function, NO2 dry deposition oh
galvanized steel can be ignored.
In a test with an NO2 and propylene mixture, Edney et al. (1986) simulated smog
conditions that might be similar to Southern California conditions (i.e., smog with very low
SO2 concentrations). This experiment was allowed to proceed in the smog chamber for
336 hours (2 weeks) with a total time of induced dew of 196 hours in 7-h periods. At the
end of the experiment, concentrations in the gas phase and in dew on the surface of the
galvanized steel were measured. Results are shown in Table 12-1. Fairly small amounts of
nitrite ions (NO2~) and nitrate ions (NO3~) were found on the surface and relatively little zinc
was freed (corroded). Clearly, however, the NO2 and other reactants had reacted to form a
number of species. >
A test with NO2, C3H3, and SO2 was also run for comparison. After 25 h, with a total
time of wetness of 14 h for the galvanized steel, the gas and surface-dew concentrations
shown in Table 12-2 were measured. The gaseous species concentrations were similar to
those found in the previous test, except for SO2. Again, little nitrate or nitrite was found in
the dew on the surface of the galvanized steel, especially when compared to the SO
X.
deposition. Furthermore, far more zinc was found in solution (i.e., corroded) when SO2 was
added to the NO2-C3H3 mixture.
August 1991
12-4 DRAFT - DO NOT QUOTE OR CITE
-------
TABLE 12-1. SMOG CHAMBER REACTIONS OF NO2 AND C3H3
AND DEPOSITION OF REACTION PRODUCTS
ON GALVANIZED STEEL
Chemical
Species
03
CH3CHO
HCHO
PAN
NOY-PAN
A
HNO3
N02-
N03-
S04=
Zn
Gas-Phase
Concentration, ppb
134
254
621
57
359
7
—
—
.
—
Surface-dew
Concentration,
nmol/cm2
—
—
133
—
—
—
11
77
133
77,
Source: Edney et al. (1986).
1 The above laboratory studies illustrate both the complex nature of the NOX chemistry
2 and the relatively low deposition rate of NOX on galvanized steel. In a subsequent field
3 experiment, Edney et al. (1987) measured the ion concentrations for dry deposition and in
4 rainwater runoff from galvanized steel samples exposed outdoors in Research Triangle Park,
5 NC. The dry deposition ratio of SO4= to NO3" was 3.4, again illustrating the relatively low
6 deposition velocity of NOX compared to SOX for galvanized steel, this time under outdoor
7 exposure conditions. These researchers speculated that the NO3~ resulted from dry deposition
8 of HNO3 and particulate nitrate. The ratio of dry to total nitrate deposition was 0.46,
9 suggesting that wet and dry deposition appeared to play abput equal roles in nitrate
10 deposition. Regression analysis of the ion concentration showed that the NO3" did not
11 significantly relate to the zinc in solution concentrations; however, SO4= concentrations were
12 ,, in a one-to-one relationship with dissolved zinc. Edney et al. (1987) concluded that NOX is
August 1991
12-5 DRAFT - DO NOT QUOTE OR CITE
-------
TABLE 12-2. SMOG CHAMBER REACTIONS OF NO2, C3H3,
AND SO2 AND DEPOSITION OF REACTION PRODUCTS
ON GALVANIZED STEEL
Chemical
Species
03
HCHO
PAN
NOX-PAN
HNO3
S02
N02-
S03=
NO3-
S04°
Zn
Gas-Phase
Concentration, ppb
240
1,150
114
159
9
1,190
—
—
—
—
—
Surface-dew ;
Concentration,
nmol/cm2
.
560
—
—
, •
—
4
595
19
91
441
Source: Edney et al. (1986).
1 not effectively deposited on galvanized steel surfaces and that sulfates dominate galvanized
2 steel corrosion.
3 While NOX deposition to galvanized steel may be insignificant, Spicer et al. (1987)
4 found that there is a significant range of removal rates of NO2 by common indoor materials.
5 Samples of 35 materials (surface area 3.3 m2) were exposed in chambers to 282 /*g/m3
6 (0.15 parts per million [ppm]) NO2 (initial condition) at 50% relative humidity (RH) for 12 h
7 and the rate of NO2 removal was measured. The results of these experiments are shown in
8 Figure 12-1. Galvanized metal ducts, probably of material similar to the metal used by
9 Edney et al. (1987), were near the low end of removal rates measured in the Spicer et al.
10 (1987) experiments. Many common indoor materials (wallboard, wool carpet) were found to
11 have very high removal rates. Nitric oxide gaseous concentrations were also monitored
12 during these experiments and were often found to increase as NO2 levels decreased. The
August 1991
12-6 DRAFT - DO NOT QUOTE OR CITE
-------
01234567
WALLBOARD
CEMENT BLOCK
WOOLCARPET
BRICK (USED)
MASONITE
COTTON/POLYESTER BEDSPREAD
PAINTED (FLAT LATEX) WALLBOARD
PLYWOOD
ACRYLIC FIBER CARPET
NYLON CARPET
VINYL WALL COVERING (PAPERBACKED)
CEILING TILE
POLYESTER CARPET
ACRYLIC CARPET
FURNACE FILTERS (NEW)
DEHUMIDIFIER
OAK PANELING
VINYL-COATED WALLPAPER
PARTICLE BOARD
FURNACE FILTERS (USED)
CERAMIC TILE
WOOL(80%) POLYESTER(20%) FABRIC
COTTON TERRYCLOTH
SPIDER PLANTS (WITH SOIL COVERED)
WALLTEX COVERING
WAXED ASPHALT TILES
WINDOW GLASS
USED FURNACE HEAT EXCHANGER
FORMICA COUNTER TOP
POLYETHYLENE SHEET
ASPHALT FLOOR TILES
VINYL FLOOR TILE
GALVANIZED METAL DUCT
PLASTIC STORM WINDOWS
01234567 8 9
RATE CONSTANT FOR N02 REMOVAL, 1/hr
Figure 12-1. Bar graph of NO2 removal rate for various materials evaluated in a
1.64 m3 test chamber at 50% RH.
Source: Spicer et al. (1987).
August 1991
12-7 DRAFT -DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
author suggested that judicious selection of indoor materials might be considered as a means
of indoor NO2 control. However, it was not possible from these experiments to determine
the amount of NOX accumulating on the surfaces of these materials, nor could conclusions be
drawn on any damage to indoor materials that might result from exposure to NO2.
MiyazaM (1984) conducted a similar experiment exposing common interior materials in
a chamber to initial concentrations of 1,645 mg/m3 (875 ppm) NO2 and 1,124 mg/m3
(914 ppm) NO. A summary of these results is shown in Table 12-3. The trend in these data
is similar to that reported by Spicer et al. (1987), with carpeting and cement showing
relatively high deposition velocities for NO2. Vinyl floor tile, glass, and metals showed
relatively low deposition velocities for NO2. Insulation board and an ester/acrylic carpet,
materials not tested by Spicer et al. (1987), had the highest deposition velocities. MiyazaM
(1984) also found that NO2 deposition rates increased if turbulence, humidity, and
temperature were each increased in the chamber. Increased turbulence escalates the rate of
delivery of NO2 to the surface. Increased humidity probably results in dissolution of NO2.
Increased temperature causes faster reaction rates.
The deposition rates reported by Miyazaki appear to be low compared to the rates
reported by Edney et al. (1986); The reason for the discrepancy is not apparent; however,
the differences may have been caused by different levels of turbulence in the two
experimental chambers. Caution should be used in applying data from Miyazaki (1984) for
more than comparative purposes.
12.1.4 Chemical Interactions of Nitrogen Oxide Species
Not only is there wide variation in the deposition of NOY to different surfaces but NO
&• X
species themselves are reactive and their interactions with other atmospheric constituents are
complex. Bassett and Seinfeld (1983) proposed a chemical equilibrium model for the
behavior of NOX, SOX, ammonium (NH3), and water in the atmosphere that is instructive for
understanding the role of NOX in materials damage. Nitrogen species (NO, NO2, HNO3,
etc.) are present as gases and in particulates (liquid and solid) and.are deposited on material
surfaces. Nitric acid is potentially the NOX species most directly damaging to materials and
is formed by photochemical reactions involving NOX in the atmosphere. Under dry
conditions, HNO3 can deposit on a surface and can cause direct damage. If liquid water is
August 1991
12-8 DRAFT - DO NOT QUOTE OR CITE
-------
TABLE 12-3. DEPOSITION VELOCITIES OF NO2 AND NO
FOR INTERIOR MATERIALS
~~~~ Deposition Velocity,
cm/s
Interior Material
Flooring materials
Carpet 1 (Acrylic Fiber)
Carpet 2 (Acrylic Fiber)
Carpet 3 (Acrylic Fiber)
Carpet 4 (Wool)
Carpet 5 (30% Ester, 70% Acrylic Fiber)
Tatami facing
Needle punch
Bath mat (100% Cotton)
Floor sheet 1 (Vinyl chloride)
Floor sheet 2 (Vinyl chloride)
Floor sheet 3 (Vinyl chloride)
Plastic tile
Ceramic tile •
N02
0.03
0.02
0.02
0.06
0.10'
0.01
0.01
0.05
0.001
0.003
0.003
0.003
0.004
NO
0.0003
—
—
—
0.003
0.0008
—
0.00
—
—
'• ___
—
Wall materials
Wallpaper 1
Wallpaper 2.
Printed plywood
Ceiling materials
Insulation board
Painted insulation board
Plaster board
Wooden cement board
Asbestos cement board
Fittings
Glass
, Painted stainless steel
Painted wood
Curtain
Fusuma paper
Shoji paper
0.002
0.002
0.001
0.11
0.06
0.02
0.03
0.04
0.00
0.0008
0.003
0.0008
0.003
0.0003
0.00
0.00
0.001
0.003
0.003
0.0008
0.001
0.0003
0.0003
0.002
0.0003
These values were averaged from the results of the experiments at 20-26 °C, 40-60% RH.
Source: Modified from Miyazaki (1984).
August 1991
12-9 DRAFT - DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
present, HNO3 exists in equilibrium between the liquid phase in water solution and the
gaseous phase in the atmosphere. However, Bassett and Seinfeld (1983) showed that in the
presence of NH3 and sulfuric acid (H.2SO£, constituents found in the atmosphere, the HNO3
gas-phase versus liquid-phase equilibrium is shifted toward the gas phase. Thus, as nitrates
accumulate on the surface of a material much of the accumulated nitrate mass can be
evaporated into the atmosphere as HNO3. Baedecker et al. (1990) believe that this
mechanism explains why most post-facto microanalytical investigations of damaged surfaces
reveal very small amounts of nitrogen species, whereas SO4= are frequently present. Thus,
in polluted atmospheres containing SO2 .and condensing moisture, it is possible that NOX
currently plays a relatively small role compared to SO2 in causing the observed damage to
most materials.
12.1.5 Materials Damage Experimental Techniques
Because of the number of possible damaging agents and the complexity of synergistic
interactions, deposition processes, and exposure scenarios, researchers have typically relied
on controlled environmental chambers to quantify the damage rates attributable to specific
agents such as NOX. Often materials exposure chamber studies are conducted at high
concentrations or at elevated temperatures and humidities in order to see damage within a
reasonable exposure period. In addition, some chamber studies are conducted at low flow
rates that poorly simulate mass transfer properties in the natural environment and lead to
underestimation of real-world deposition rates. Also, the sequence with which materials are
exposed to different pollutants can affect the formation of protective corrosion films, and this
process is sometimes poorly simulated in chambers. While such studies are useful, care
should be exercised in the extrapolation of data and conclusions based on chamber studies to
ambient exposures.
The alternative to chamber studies has been ambient exposure studies. In these
exposure studies, the materials of interest are usually exposed to ambient conditions at several
locations representing a spectrum of environmental variables (e.g., temperature, sunshine,
humidity, pollutant concentrations). Statistical and chemical analyses are then used to assess
the contribution of the measured environmental variables to the materials damage. Again, the
number of possible agents and the complexity of synergistic interactions makes it difficult to
August 1991
12-10 DRAFT - DO NOT QUOTE OR CITE
-------
1 apportion observed damage among all the possible causes. Franey and Graedel (1985)
2 reviewed the pollutant species that induce damage under actual ambient exposure conditions,
3 and have suggested that for any chamber study to be realistic, moisture, radiation, carbon
4 dioxide (CO^, reduced sulfur, a chlorine-containing gas, and a nitrogen-containing gas must
5 be included. Because of the difficulties involved in apportioning the causes of damages,
6 reliable appraisals of the damage induced by NOX exposure alone are not yet available.
7 Both chamber studies and ambient exposure studies have come to rely on sophisticated
8 ' surfa'ce chemistry analytical techniques, as well as traditional bulk chemistry analyses and
9 measurements of physical properties. Additionally, moisture collected from the samples (run-
10' off) has been analyzed for its chemical constituents. The objectives of these efforts are to
11 understand the chemical reactions occurring on the sample surfaces.
12 Generally, little evidence of NOX species has been found in these analyses. As noted in
13 the previous section, much of the NOX will be converted into HNO3 and subsequently will be
14 evaporated back into the atmosphere. Thus, if HNO3 is leading to damage, it may not be
15 adequately accounted for in either surface chemical or runoff chemical analyses; and its role
16 in the damage process could be underestimated. Better experimental techniques are needed
17 both for investigating materials damage on the whole, and for determining the role played by
18 NOX.
19
20
21 12.2 EFFECTS OF NITROGEN OXIDES ON DYES AND TEXTILES
22 12.2.1 Fading of Dyes by Nitrogen Oxides
23 Textile and dye manufacturers have recognized the problem of dye fading induced by
24 NO for some time. Rowe and Chamberlain (1937) reported that dyes fade because of the
A
25 presence of NOX in combustion effluents. Carpets, upholstery, and drapes, subjected to
26 elevated NC- levels in buildings using unvented gas heat, have been observed to fade within
X. -
27 a year when dyes not resistant to NOX fading have been used. Fading is exacerbated when
28 susceptible fabrics are dried in gas-fired clothes dryers, in which the concentrations of NO2
29 can reach 3,760 ^tg/rn3 (2.0 ppm) (McLendon and Richardson, 1965). Moreover, dryer
30 exhaust is sometimes vented to the indoor environment to conserve heat and humidity, thus
31 increasing indoor concentrations of NOX. Textile and dye manufacturers have responded to
August 1991
12-11 DRAFT - DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
NOx-induced deterioration by seeking out and using NOx-resistant dyes or inhibitors that
forestall fading. Fading from NOX has been observed on acetate, cotton, nylon, rayon, wool,
and polyester.
Nitrogen oxide-induced ("gas-fume") fading received serious attention when blue
disperse dyes were found to deteriorate significantly on cellulose acetate. Salvin and
coworkers (1952) pointed out that NO2 is soluble in cellulose acetate, and that during
laboratory tests significant fading of dyes on the material can be observed within an hour.
Hemphill et al. (1976) tested a spectrum of dyes on various fabrics and found that NO2
caused significant fading on the cellulose acetate samples. Salvin and Walker (1959) and
Salvin (1964) showed that alternative dyeing processes are available to minimize the impact
of NOx-induced fading on cellulose acetate, but that in many cases these substitute processes
and dyes are more expensive to use than the processes and dyes they replaced.
Beloin (1973) exposed a variety of fabrics and dyes to 120 )Wg/m3 (0.1 ppm) and
1,230 /ig/m3 (1 ppm) of NO, and 90 jitg/m3 (0.05 ppm) and 940 ^g/m3 (0.5 ppm) of NO2
for 12 weeks in an environmental exposure chamber. He found that "appreciable" to "very
much" (the most severe category) fading occurred at both concentrations of NO for cottons
with direct, reactive, and vat blue dyes, cellulose acetate with disperse blue dyes, and nylon
with a blue dye. The cellulose acetate samples exposed to NO2 had generally greater
amounts of color change than the samples exposed to NO. In addition, NO2 affected cotton
with direct and reactive red dyes, cotton with reactive blue dye, and rayon with direct red
dye. Beloin (1972) conducted tests on 67 dye-fabric combinations at 11 urban and rural sites
nationwide for 3-month exposures. The tests were conducted outdoors using chambers
designed to let the ambient air circulate around the samples but to exclude sunlight. Using
multiple regression analysis, he sought to determine which pollutants played a significant role
in the observed change of colors on the fabrics. He found that SO2 concentrations were
significant for 23 fabrics, ozone (O3) was significant for 8 fabrics, and NO2 was significant
for 7 fabrics. Fabric-dye combinations affected by NO2 included cellulose acetate with red
and blue disperse dyes, cotton muslin with reactive red and blue dyes, wool flannel with acid
blue dye, and the NOX gas-fading control ribbon recommended by the American Association
of Textile Chemists and Colorists for testing NOX fading.
August 1991
12-12 DRAFT - DO NOT QUOTE OR CITE
-------
1 Cotton is the most widely used natural textile fiber and, again, significant gas-fume
2 fading has been noted. Haynie et al. (1976) exposed plum-colored cotton drapery fabric to
3 NO2 in a chamber for 1,000 h and found that serious fading occurred. Based on
o
4 extrapolation they predicted that the use-life of draperies exposed to 100 jKg/nr (0.053 ppm)
5 NO2 would decrease 19%. In Beloin's chamber study described above, dyes on cotton were
6 found to experience "noticeable" to "much" fading when exposed to NO and "noticeable" to
7 "very much" fading when exposed to NO2. McLendon and Richardson (1965) found that
8 blue-dyed cotton fabric became green after repeated NOX exposures in gas-fired dryers and
9 that the NOX exposure caused white fabric to "yellow". Salvin (1969) reported the results of
10 „ sheltered, outdoor exposures of dyed cottons for 90 days in Los Angeles. Thirty-one colors
11 of direct, vat, reactive, and sulfur dyes were tested and fifteen faded substantially. The
12 author concluded that NOX and. O3 were primarily responsible. Hemphill et al. (1976) also
13 demonstrated NOx-induced fading of vat, direct, and reactive dyes on cotton at concentrations
14 of 940 Aig/m3 (0.5 ppm) in a chamber for a 5-h exposure. -
15 Imperial Chemical Industries Limited, (1973), a supplier of dyes for synthetics, issued a
16 technical bulletin on the gas-fume fastness of dyes used for nylon (polyamide). Nylons have
17 high resistance to wear and thus are often used as carpeting. In this application nylons are
18 exposed to indoor atmospheres for long periods. Imperial Chemical Industry's bulletin
19 , showed that several of the commercially available dyes faded noticeably on nylon when
20 exposed to NOX fumes and advised that these dyes not be used. The susceptible dyes fade,
21 become duller in appearance, or acquire a redder or yellower cast. Hemphill et al. (1976)
22 demonstrated that certain blue and red dyes on nylon fade substantially when exposed to
23 940 jug/m3 (0.5 ppm) NO2. Beloin's (1973) chamber study found that "appreciable" to very
24 much" fading occurred on nylon fabrics exposed to NO or N02. In outdoor exposures in Los
25 Angeles, Salvin (1964) found that nylon faded only slightly to very slightly.
26 Other fabrics have been tested for dye gas-fading resistance as .well. Hemphill et al.
27 (1976) investigated dye fading of rayon. They found that two of the dyes tested^ Direct Blue
28 .• 86 and Direct Red 79, showed '.'noticeable" to "significant" fading. Beloin (1973) found that
29 rayon withstood NO exposure with only a trace of fading, but exhibited "very much" fading
30 when exposed to NO2. In checking orlon, Hemphill et al. (1976) found minimal dye fading.
31 Salvin (1964) found that wool did not fade significantly in Los Angeles ambient exposures,
August 1991
12-13 DRAFT - DO NOT QUOTE OR CITE
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
but Hemphill et al. (1976) showed moderate fading of red dye on wool in chamber
exposures. Polyester exhibited very good dye-fading resistance in Salvin's Los Angeles study
(1964).
Hie American Association of Textile Chemists and Colorist encourages textile
manufacturers and suppliers to test dye and fabric combinations for NOX fading. These tests
are routinely performed and NOx-susceptible dye and fabric combinations rarely are produced
in large quantities for the open market.
12.2.2 Degradation of Textile Fibers by Nitrogen Oxides
Nitrogen oxides not only affect fabric color, but can also alter the physical
characteristics of the fiber themselves. Jellinek (1970) and Jellinek et al. (1969) reported
significant chain scissioning of nylon after NO2 exposure. Chain scissioning is the breaking
of the molecular structure that makes up a polymer and it results in a loss of strength.
Vijayakumar et al. (1989) found statistically significant amounts of damage to nylon textiles
exposed for 28 days to 0.1 ppm and 0.5 ppm concentrations of HNO3. Zeronian et al.
(1971) investigated the impact of NO2 on acrylic, modacrylic, nylon, and polyester yarn.
The yarns were continuously exposed in chambers for 1 week to simulated sunlight and
3,760 /ig/m (0.2 ppm) NO2. The yarn strength and rupture energies were reduced for all
materials. The most seriously affected was nylon yarn, which lost approximately 30% of its
strength and 33% of its rupture energy as compared to control samples exposed without NO2.
The least affected was polyester, with about a 13% decrease in strength. The loss of strength
of the acrylics was intermediate between the other two yarns.
12.3 EFFECTS OF NITROGEN OXIDES ON PLASTICS AND
ELASTOMERS
12.3.1 Chemical Changes Induced by Nitrogen Oxides
Plastics are highly polymerized materials, mostly synthetics, combined with other
constituents such as hardeners, fillers, and reinforcing agents (Hawley, 1981). Plastics
include fluorocarbon resins, phenolics, polyimides, polyethylene, acrylic polymers,
polystyrene, polyurethane, and numerous other synthetic compounds. Major uses of plastics
August 1991
12-14 DRAFT - DO NOT QUOTE OR CITE
-------
1 include automobile bodies and components, boat hulls, building and construction materials
2 (pipe, siding, flooring), packaging (bottles, vapor barriers, drum linings), textiles (carpets,
3 cordage, hosiery), organic coatings such as paint and varnish vehicles, adhesives, electrical
4 components, and numerous other applications. Use of plastics in the United States in 1980
5 was estimated at approximately 60 billion pounds per year, or double the 1970 consumption.
6 Further development of arid reliance on plastics are expected to increase the demand for them
7 in the future. Elastomers are synthetic polymers with the ability to stretch to at least twice
8 their normal length and retract rapidly to near their normal length when released. Examples
9 of elastomers include butyl, nitrile, and polysulfide rubber, and neoprene. Elastomers are
10 used for vibration dampers, wire coatings, fabrics, automobile tires, bumpers and windshield
11 wipers, and other applications.
12 Plastics and elastomers are subject to deterioration on exposure to ultraviolet radiation
13 (UV), O3, SO2, and NOX. Jellinek et ai. (1969) and Jellinek (1970) reported a series of
14 experiments in which a variety of polymers and elastomers were exposed to radiation and
15 pollutants in chamber experiments. Jellinek et al. (1969) reported the following results for
16 , high concentration (nearly pure) NO2 exposures.
17
Polyethylene: minimal effect except for an increase in viscosity.
Polypropylene: some cross-linking (forming of additional chemical bonds) of the
polymer, although not as much as when exposed to SO2.
Polystyrene: some chain-scissioning (breaking of chemical bonds).
Polymethyl methacrylate: some chain-scissioning (breaking of chemical bonds).
Polyvinyl chloride: loss of chlorine due to reaction with NO2. <
Polyacrylonitrile: no significant change.
Nylon: chain-scissioning occurs. ,
Butyl rubber: chain-scissioning. ,
Polyisoprene: appreciable chain-scissioning.
Polybutadiene: cross-linking'occurs. ,
18
19
20
21
22
23
24
25
26 '•
27
28
29
30
31
32
33
34
35
36
37
38
1.
2.
3.
4.
5.
6.
7.
8.
9.
10.
August 1991
12-15 DRAFT - DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
They concluded that damage to elastomers was generally greater than damage to plastics, but
that O3-induced damage was probably more important than NO2-induced degradation.
Jellinek (1970) reported findings for the same series of plastics and elastomers at NO2
concentrations of 1,880 jttg/m3 and 9,400 /ig/m3 (1 and 5 ppm) for 1 h exposures. At these
levels polymethyl methacrylate, nylon, and butyl rubber were found to suffer chain-
scissioning. Polyethylene, polypropylene, polyisoprene, and polybutadiene exhibited cross-
linking. ,
Krause et al. (1989) exposed polyvinyl chloride, polyurethane, glass-fiber-reinforced
polyester, and alkyd resin for 5 years in glass chambers to either 5,000 jttg/m3 NOX,
5,000 ng/m3 SO2, 2,500 jttg/m3 O3, or a mixture of the pollutants. The exposure cells were
kept at a humidity of 50 to 60%. Half of each chamber was exposed to sunlight through
UV-transmitting glass. The other half was kept dark. The investigators found that most of
the degradation was caused by sunlight, with significantly less degradation occurring from
dark exposures to pollutants.
Haynie et al. (1976) exposed tire rubber and vinyl house siding to NO2, SO2, O3,
radiation, and humidity in a chamber. Two NO2 concentrations, 94 and 940 jiig/m3
(0.05 and 0.5 ppm), Were used with exposure times of 250, 500, and 1,000 h. Various
combinations of exposures with the other pollutants, radiation, and humidity conditions were
used. The primary cause of damage to rubber was O3 exposure and NO2 actually seemed to
inhibit the rate of O3-induced damage. No appreciable damage to vinyl siding was observed.
The National Research Council (1977) notes that discoloration and deterioration of strength of
foam rubber occurs from NO2 exposure.
12.4 EFFECTS OF NITROGEN OXIDES ON METALS
12.4.1 Role of Nitrogen Oxides in the Corrosion Process
Atmospheric corrosion of metals is a serious problem and air pollution is known to
accelerate corrosion processes. Sulfur oxides and chlorides are the atmospheric contaminants
most frequently implicated in the corrosion of metals. Nitrogen oxides are also involved but
have received less attention. Moisture enables these contaminants to form aggressive acids
that attack the metal surface and promote electrochemical reactions. For this reason, both
August 1991
12-16 DRAFT - DO NOT QUOTE OR CITE
-------
1 pollutant concentrations and the "time of wetness" (i.e., the duration for which the material
2 surface has liquid water present) for exposed surfaces are important in determining the
3 amount of damage that will occur.
4 For most metals, NQX alone as an attacking agent is much less aggressive than sulfur or
5 chlorine compounds. Svedung et al. (1983), Kucera (1986), and Johansson (1986), however,
6 . have pointed out the synergistic impact of NOX or/atmospheric corrosion mechanisms. Using
7 an exposure chamber, Kucera (1986) showed that carbon steel corrodes rapidly when exposed
8 to:3,421 )ug/m3 SO2 and 90% RH, but very slowly when exposed to SO2 at the same
9 concentration and 50% RH. At 50% RH humidity, steel corrodes about three times more
10 quickly when exposed to NO2 (5,640 jug/m3). However, when both NO2 and SO2 at the
11 same concentrations are present at 50% RH, the .corrosion rate is approximately 30 times the
12 rate seen with SO2 alone. Kucera noted that the presence of NO2 increases the rate of
13 deposition of SO2 on the metal surface. Johansson (1986),. also using an exposure chamber,
14 showed that NO2 deposition leads to the formation of hygroscopic nitrate-containing corrosion
15 products on the surface of the metal. These corrosion products, in turn, absorb moisture onto
16 the surface, making the moisture available to mobilize other ions (such as sulfates and
17 chlorides) and thus leading to active corrosion at much lower relative humidities than if NO2
18 were, not present. Effectively, NO2 acts to increase the time of wetness for the surfaces.
19 Svedung et al. (1983) showed similar results for gold-coated brass (a common electrical
20 contact), with NO2-containing atmospheres accelerating degradation at all humidity levels
21 between 40 and 80%.
22 In the outdoor environment, the deposition of NO2 is limited, for most materials, by the
23 surface uptake resistance; and NO2 is more slowly adsorbed than SO2. In the experiments
24 conducted by Svedung et al. (1983), Kucera (1986), and Johansson (1986), low flow rates
25 were used in the exposure chambers. During low flow conditions, the deposition rate
26 becomes limited by the surface boundary layer resistance and the effective deposition rates of
27 NO2 and SO2 may become nearly equal. Thus, the conclusion that NOX is synergistic with
28 SO2 may not be applicable in outdoor environments. In indoor exposures of materials,
29 however, the conclusions of Svedung et al., Kucera, and Johansson may be applicable.
30
31 ... .-,,..
August 1991
12-17 DRAFT - DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
12.4.2 Effect of Nitrogen Oxides on Economically Important Metals
Steel
Steel is the most widely used structural metal and is available in a wide variety pf types
with varying percentages of alloying elements. Basically, steel consists of iron containing
0.02 to 1.5% carbon. The corrosion behavior of common construction steels (carbon steels,
containing about 0.2% carbon) is similar, and rusting of exposed surfaces proceeds rapidly.
Low alloy steels, containing chromium, nickel, copper, molybdenum, phosphorus, and
vanadium in the range of a few percent or less for the total inclusion, are substantially
stronger and offer improved resistance to atmospheric corrosion. Specialty steels, such as
stainless steels containing over 10% chromium, are designed to be highly corrosion-resistant,
but are also much more costly. Bare steel is not usually exposed to the environment, but
rather is painted to prevent rust and premature failure. Nevertheless, except where
specifically noted, the following discussion concerns common construction steel that is boldly
exposed with no coatings.
Samples of enameling steel were exposed at 57 of the National Air Site Network
locations (Haynie and Upham, 1974), for 1- and 2-year exposure cycles. Sulfur dioxide and
particulate matter concentrations, relative humidity, and paniculate chemistry were monitored
at the sites. Corrosion rates for the steel samples were determined from weight loss
measurements, and these data were correlated against the pollution measurements. Haynie
and Upham (1974) concluded that either SO2 or particulate sulfate, or both, were significant
in causing steel corrosion. Particulate nitrate (PN) was not statistically significantly related to
the observed corrosion; however, measurement techniques for PN were unreliable.
Measurements of gaseous NOX species were not made.
Johansson (1986) showed in a low flow chamber study that gaseous NO2 adsorbs on
steel surfaces and reacts with water to form HNO3 and nitrous acid (HNO2). Construction
Steel was exposed continuously for 6 weeks to 376 ^cg/m3 or 5,640 /-tg/m3 (0.2 or 3.0 ppm)
NO2 and different levels of moisture and SO2. He determined that the deposition rate of
NO2 was much lower than the deposition rate for SO2 and that, if no other pollutants were
available, steel exposed to NO2 alone will slowly acquire a thin oxide layer (rust) that
protects the underlying steel from further damage. Unfortunately, the nitrates formed during
the corrosion process are hygroscopic and act to adsorb further moisture from the atmosphere
August 1991
12-18 DRAFT - DO NOT QUOTE OR CITE
-------
1 at relative humidities around 50% and above. If it is also present, SO2, which does not form
2 hygroscopic corrosion products but does have a higher deposition rate than NO2 (Johansson,
3 1986), reacts with this moisture to form strong acids that corrode the surface very rapidly. In
4 addition to its hygroscopic effect, Johansson suggested that NO2 might increase the oxidation
5 rate of SO2 to SO4=, and thus enhance corrosion. At relative humidities in excess of 90%,
6 the synergistic effect of NO2 is lost because at these high humidity levels moisture forms on
7 the surface whether or not NO2 is present. In fact, Henriksen and Rode (1986) have
8 suggested that NO2 may actually inhibit SO2-induced steel corrosion at relative humidities of
9 95%.
10 Haynie (1986) analyzed data from 30 months of exposures of weathering steels at nine
11 sites around St. Louis, MO, as part of the U.S. Environmental Protection Agency's (EPA's)
12 Regional Air Pollution Study (RAPS). Weathering steels are architectural steels specifically
13 formulated to rapidly develop a surface corrosion layer that protects the underlying substrate
14 steel. The exposure samples were collated with air quality monitoring stations. Haynie
15 (1986) statistically analyzed the observed corrosion versus meteorological and air quality
16 variables. He found that the sample weight change was positively correlated with the SO2
17 levels, but negatively correlated with NO2. He concluded that NO2 decreases the solubility
18 of the corrosion layer.
19 Haynie et al. (1976) studied weathering steel in an exposure chamber. While they
20 concluded that NO2 did not have as significant an impact as SO2 on the indicated corrosion, a
21 review of the data showed that at low relative humidities the samples showed somewhat more
22 damage at high NO2 concentrations (940 jtig/m3, 0.5 ppm) than at low concentrations
23 (94 jtig/m3, 0.05 ppm).
24
25 Galvanized Steel and Zinc
26 Because most carbon steels rust readily when exposed to moist air, a layer of zinc is
27 frequently coated or galvanized onto the surface. The zinc acts to protect the substrate steel
28 electrochemically by preferentially corroding away, leaving the steel rust-free. Zinc
29 galvanized steel is used for many outdoor purposes including chain-link fences, highway
30 guard rails and sign posts, roofing, and automobile body panels.
August 1991
12-19 DRAFT - DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
Whitbeck and Jones (1987) studied the accumulation of nitrates on galvanized steel in an
exposure chamber. They exposed the galvanized steel to 18,800 /*g/m3 (10 ppm) of NO2 and
measured the nitrate formation as a function of time on the sample surface. They found that
the formation of nitrates was linear with time. Haynie et al. (1976) included galvanized steel
in their chamber study discussed above and concluded that the effects of SO2 are much more
significant than those of NO2.
These results are further supported by the field investigations reported by Cramer et al.
(1988). They found that SO2 is more readily absorbed on galvanized surfaces than NO and
NO2 and that SO2-induced corrosion probably dominates corrosion by NOX in most
environments. In relatively dry environments, Cramer et al. (citing Johansson, 1986) pointed
out that NO2 can participate in a reaction to oxidize SO2 and form sulfuric acid (H2SO4),
which is very aggressive to galvanized surfaces. Edney et al. (1987) statistically analyzed the
results of exposures of galvanized steel and chemical analyses of the runoff rainwater from
the samples. They found that me amount of deposited SO4= dominated the amount of
deposited NO3-, and that SO4= and NO3~ deposition rates were strongly correlated at the field
exposure site. The regression analysis, therefore, found that SO4= dominated the corrosion
of galvanized steel and that NO3' was not a significant contributor to corrosion at this
location. Subsequent analysis of data from the same site by Spence etal. (1988), using a
more complete regression model, found no statistically significant effects of pollution on
either galvanized steel or weathering steel exposed for 3 years. The site used for this
experiment, Research Triangle Park, NC, is relatively rural and SO2 and NO2 concentrations
are fairly low. The analysis of Spence et al. suggests that natural weathering processes
dominate over corrosion at this site.
Although rarely used alone as a construction material, zinc is used for galvanizing and
as an alloying metal and its corrosion behavior has been investigated. Johansson (1986)
exposed zinc to NO2 and SO2 in a low-flow exposure chamber. He showed that NO2 alone
had little impact, but was strongly synergistic when combined with SO0. As the NOo
•^ i
concentration in the mixture was increased from 376 ,«g/m3 to 5,640 j«g/m3 (0.2 ppm to
3.0 ppm), and the SO2 concentrations were held constant; there was little change in the rate
of corrosion.
August 1991
12-20 DRAFT - DO NOT QUOTE OR CITE
-------
1 Kucera (1986) has noted that, in the open air, zinc tends to form a layer of sulfates and
2 carbonates on the surface that acts to passivate the metal. This layer is basic; and if rain with.
3 a pH value of 4 or less washes the surface, the layer is removed, exposing the substrate
4 metal. In this way zinc is sensitive to acid deposition, so that any pollutant, including NOX,
5 that adds to the acidity of the environment is damaging to zinc.
6 Hermance (1966) and Hermance et al. (1971) reported the impact of nitrates on zinc-
7 *. containing nickel-brass wire springs used in telephone relays. They pointed out that
8 hygroscopic nitrate salts collected on the springs and moisture formed on the surface at any
9 relative humidity exceeding 50%. The nitrate deposition resulted in attack on the zinc in the
10 springs and premature failure of the relays. In addition, Graedel and McGill (1986) have
11 .pointed out that NO2 is known to be moderately aggressive towards nickel. Ultimately, the
12 telephone companies were forced to replace zinc-containing nickel-brass springs in areas with
13 high NOX levels, such as Los Angeles. Henrikson and Rode (1986) showed that at 95% RH
14 the synergistic effects of NO2 and SO2 were not detectable for zinc corrosion. At high
15 humidities SO2 appears to dominate zinc corrosion.
16
17 Aluminum
18 ' Aluminum is widely used because of its corrosion resistance and is second only to steel
19 in the amount of metal in use. Aluminum is often exposed without coatings, such as paint,
20 and is used for architectural trim, aircraft, small buildings, cooking utensils, etc. Kucera
21 (1986) noted that the time of wetness of aluminum surfaces correlates with NOX
22 concentrations, but could not conclude that NOX was of any practical importance in the
23 aluminum corrosion process. Johansson (1986) demonstrated in a chamber study that NO2
24 did not significantly adsorb on aluminum but that at 90% RH NO2 was synergistic with SO2
25 and caused nearly three times the corrosion caused by either pollutant alone. Henriksen and
26 Rode (1986) showed that NO2 inhibits SO2-aluminum corrosion at 95% RH. In a chamber
27 study, Loskutov et al. (1982) demonstrated that the interaction of NO2 and water on an
28 aluminum surface was a complex process. They concluded that adsorbed water acted to
29 displace NOX on the surface; and that metal corrosion occurred simultaneously with the
30 adsorption/displacement process but slowed substantially as water displaced NOX.
August 1991
12-21 DRAFT - DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
I' 20
21
22
23
24
25
26
27
28
29
30
31
Vijayakumar et al. (1989) exposed aluminum to 940 and 1,880 jwg/m3 (0.5 and 1 ppm)
NO2 in a chamber for 28 days. They found no statistically significant impact of NO2 on
aluminum. They also exposed aluminum to 252 and 1,260 jug/m3 (0.1 and 0.5 ppm) HNO3
and determined that there was statistically significant damage and that the rate of the
damaging reaction was relatively rapid.
Copper
Copper is used for architectural trim, electrical components, and heat transfer coils in
air conditioners. Chamber studies (Schubert, 1978; Rice et al., 1981) have shown that NO2
has little impact on copper at concentrations up to 2,444 /*g/m3 (1.3 ppm). Rice et al.
(1980a) concluded from a multiple-city exposure study that hydrogen sulfide (H2S), SO2, and
O3 all had more impact than NOX on copper. Kucera (1986), Johansson (1986), and
Henriksen and Rode (1986), using chamber studies, found that the NO2 and SO2 combined
was synergistic and increased the observed corrosion rate of copper by ten to twenty times the
rate observed with single gas exposures under low-flow-rate conditions.
Nickel
Nickel is used as a coating material to protect other metals from corrosion and is
particularly resistant to environments that aggressively attack steels, aluminums, and a variety
of other metals (e.g., marine environments). Rice et al. (1980a) investigated the indoor
corrosion of nickel in several urban areas and found that SO2, NO2, and chlorides played a
significant role in accelerating nickel corrosion. In a chamber study, Rice et al. (1980b)
found that NO2 attacked nickel but that SO2 and chlorine (C12) were more aggressive than
NO2. Graedel and McGill (1986) have listed NO2 as being moderately aggressive toward
nickel.
12.4.3 Effects of Nitrogen Oxides on Electronics
While the impact of air pollution on architectural and structural metals in the outdoor
environment has been recognized for some time, the attack of NOX on electronic components,
generally used in indoor environments, is a more recently recognized problem. Telephone
companies first reported the problem with failures of wire-spring relays in telephone
August 1991
12-22 DRAFT - DO NOT QUOTE OR CITE
-------
1 switching offices located in regions with high NOX levels (Hermance, 1966; McKinney and
2 Hermance, 1967; Hermance et al., 1971). Nitrogen oxides were depositing on the springs
3 and eventually leading to stress corrosion failures. Here, the cost of the failed part, the,
4 spring, was a minor consideration compared to the loss of service. Eventually, technology
5 made the wire-spring relays obsolete, but meanwhile inconveniences and costs were incurred
6 as the result of these failures.
7 Most of the gold used for industrial purposes is used to inhibit corrosion in electrical
8 contacts. Svedung et al. (1983) tested the corrosion resistance of gold-plated brass, one of
9 . the most common contact materials, in an atmosphere containing 940 ^g/m3 (0.5 ppm) NO2.
10 They found that NO2-containing environments were more aggressive than SO2 environments
11 at all relative humidities from 40 to 80%. As found with common metals, an environment
12 containing a mixture of NO2 and SO2 was even more damaging. Samples of gold contacts
13 exposed to mixed gas atmospheres became partly covered by visible corrosion after 2 to 3 h.
14 Kucera (1986) reported similar findings for electrolytic copper contacts. Buildup of corrosion
15 layers on electrical contacts causes loss of conductivity and possible failure of the contact.
16 Voytko and Guilinger (1988) exposed gold, nickel, and palladium samples electroplated
17 on copper substrates to an atmosphere containing 100 ppb NO2, 100 ppb H2S, and 10 ppb
18 Clo at 60% RH for 332 h. These samples were designed to simulate typical electrical contact
£
19, materials. They found that all coatings developed "pores" which allowed the substrate copper
20 to corrode and that the "solderability" of the specimens generally decreased after exposure.
21 Graedel and McGill (1986) reviewed the impact of pollutants on a variety of materials, and
22 listed NO2 as being moderately aggressive to solder.
23 Abbott (1987) exposed electrical contacts made of cobalt-hardened gold over sulfamate-
24 nickel to different pollutant mixtures in a laboratory test environment. He found that H2S
25 and SO2, both singly and in combination, were fairly benign to the contact surfaces, even as
26 concentrations approached 1 ppm, producing only mild pore corrosion. The reaction became
27 more severe when NO2 was added to the mixture. A mixture of 0.1 ppm H2S plus 0.1 ppm
28 SO2 plus 0.1 ppm NO2 was more aggressive than 0.5 ppm H2S plus 1.0 ppm SO2. Abbott
29 . also estimated that approximately 30% of indoor electrical and electronic equipment
30 environments are corrosive enough to result in pore corrosion and film creep that could lead
31 to component failure.
August 1991
12-23 DRAFT - DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
Freitag et al. (1980) investigated the corrosion of magnetic recording heads of the types
used in computers. They found that exposure to 0.3 ppm of NO2 and SO2 led to the
formation of corrosion products on the heads. This corrosion would lead to a degradation of
the magnetic properties of the recording head.
12.5 EFFECTS OF NITROGEN OXIDES ON PAINTS
Paints are by far the dominant class of manmade materials exposed to the atmosphere in
both indoor and outdoor environments. Paint systems are used to protect substrate materials
such as wood, steel, and stucco from damaging environmental agents, including moisture, ;
sunlight, and pollutants. Paints are also applied for aesthetic reasons. Paints are broadly
classified as architectural coatings (e.g., house paints, stains, varnishes), product coatings
(e.g., furniture finishes, automotive paints, appliance coatings), and special-purpose coatings
(e.g., bridge paints, swimming pool coatings, highway marking paint).
While paints are designed to erode uniformly and repainting is expected, any damaging
process that exposes the substrate material or discolors the finish more rapidly than natural
weathering results in premature failure of the paint system and leads to the need for
maintenance and thus to increased costs. Major paint manufacturers routinely conduct
proprietary tests of their coatings, and some information is available in the open literature
about the effects of NOX on selected paint systems. Because paint formulations vary widely,
however, results obtained for one paint may not be directly applicable to other paints.
Spence et al. (1975) investigated the effects of various pollutants on oil-based house
paint, vinyl coil coating, and acrylic coil coating. A chamber study approach was used with
1,000 h of exposure to 94 and 940 /*g/m3 (0.05 and 0.5 ppm) NO2 in combination with
various levels of SO2, O3, and humidity. The coil coatings were very resistant to all
pollutants and showed little change over the course of the experiment. The oiR>ased house
paint was found to be most sensitive to SO2 and humidity, but increased concentrations of
NO2 led to increased sample weights. This implies that the NO2 was reacting with the paint
in some way, although whether this reaction was significant was not discussed.
Haynie and Spence (1984) reported results of exposures of latex and oil exterior house
paints for 30 months at nine sites around St. Louis, MO. They reported that NO became
X
August 1991
12-24 DRAFT - DO NOT QUOTE OR CITE
-------
1 .incorporated into the latex paint film and suggested that it reacted with the polymers that
2 make up the paint. Similar results were reported for oil-based paint and brown staining.
3 Vijayakumar et al. (1989) exposed samples of high- and low-carbonate paints to NO2
4 and HNO3 for 28 days in an exposure chamber. They found statistically significant damage
5 to low-carbonate paints at 940 /xg/m3 (0.5 ppm) NO2, but not at 1,880 /*g/m3 (1 ppm) NO2.
6 The amount of damage was slight. At 1,260 jug/m3 (0.5 ppm) HNO3, however, both
7 carbonate and non-carbonate paints were damaged.
8 • • • • '
9 ; ... . • . ••.••• • • • "'
10 12.6 EFFECTS OF NITROGEN OXIDES ON STONE AND CONCRETE
11 Air pollution has been known to damage both building and statuary stone. Many
12 famous edifices, such as the Taj Mahal and the Parthenon in Athens, have been the subject of
13, studies of air pollution-induced damage to building stone. Calcareous stone, such as
14 limestone, marble, and carbonate cemented sandstone, is subject to air pollution attack.
15 Silicate stone, such as granite, slate, and non-carbonate sandstone, is much less susceptible.
16 The effects of SO2 deposition on calcareous stone are well-documented because calcium
17 sulfate (gypsum) has limited solubility and remains on protected stone surfaces as a dark
18 gypsum coating. Calcium nitrate resulting from direct NOX attack is both very soluble and
19 hygroscopic and thus washes off the stone surface almost as soon as it is exposed to rain.
20 Livingston and Baer (1983) suggest that the solubility of calcium nitrate has caused many
21 researchers to overlook NOX deposition to stone. Thus, while few data are available, NOX
22 , may have a significant effect on certain types of stone.
23 The interaction of NCL with building stone is complex. Not only will 'nitrogen
A.
24 compounds interact directly with the stone, but various endolithic bacteria present in the stone
25 result in biochemical interactions (Baumgaertner et al., 1990). Nitrosomonas spp. oxidize
26 ammonium to nitrous acid and Nitrobacter spp. oxidize HNO2 to HNO3. Production of these
27 acids results in direct chemical attack to calcareous stone and concrete. Baumgartner et al.
28 have also reported that the surface of construction stone is a significant source of NO,
29 apparently, biologically produced. On the other hand, NO2 and NH3 are absorbed by the
30 stone.
August 1991
12-25 DRAFT - DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
Baedecker et al. (1990) summarized the work of several researchers for the National
Acid Precipitation Assessment Program (NAPAP). They noted that by far the greatest
chemical erosion of calcareous stone results from the natural constituents of clean rain.
Carbon dioxide dissolved in rain forms carbonic acid that reacts with the calcium of the
stone. Baedecker et al. estimated that wet-deposited hydrogen ions from all acid species
account for about 20% of the chemical weathering of the NAPAP limestone and marble
samples. Dry deposition of SO2 was responsible for approximately 6 to 10% of the chemical
weathering; and dry deposition of HNO3 (believed to be the major form of NOX attack)
accounted for 2 to 6% of chemical erosion. They noted that an adequate model for
predicting dry deposition of HNO3 to stone is not available, and suggested that this topic
needs further research.
Mansfeld (1980) performed a statistical analysis of damage incurred on marble samples
exposed for 30 months at nine air quality monitoring sites around St. Louis, MO. He
concluded that NO3 and total suspended particulate (TSP) levels best correlated with observed
stone degradation; however, the analytical techniques used may be questionable and could
have resulted in inappropriate conclusions. Livingston (1985) reviewed current studies
regarding the impact of NOX on calcareous stone. He concluded that sulfates dominate the
damage to stone, but that NOX can play a role. Livingston also showed that the reaction of
stone with SO2 is thermodynamically favored over the reaction with NO2, and that if both
pollutants are present more calcium sulfate than calcium nitrate will be formed. Amoroso
and Fassina (1983) have suggested that the primary impact of NOX on stone may be its role
in oxidizing SO2 to form sulfate and eventually H2S04. Although this is not a direct NO
X
attack, it does lead to the degradation of stone.
Johansson et al. (1988) exposed limestone, marble, and travertine to SOo and NO for
£ X
6 weeks at various concentration combinations in the ppm and sub-ppm range. The exposure
chamber flow rates were low, with a net "wind speed" over the samples of only 0.004 m/s.
The investigators found that significantly more gypsum formation occurred with the
combinations of pollutants than with either pollutant alone. The low flow rates in the
chamber, however, make these data questionable for direct application to outdoor exposures. '
Concrete is a widely used construction material and dominates infrastructure
construction (bridges, highways, water and sewer systems). Webster and Kukacka (1985)
August 1991
12-26 DRAFT - DO NOT QUOTE OR CITE
-------
1 surveyed the construction industry and the technical literature for information regarding the
2 impact of pollutants on concrete and cement. They speculate that HNO2 and HNO3 are more
3 damaging than H2SO4 to concrete on brief exposures because they convert the Ca(OH)2 to
4 very soluble calcium nitrate. They also believe that even highly diluted HNO3 solutions can
5 bring about extensive destruction to concrete.
6
7 . ;. • . • . ."
8 12.7 EFFECTS OF NITROGEN OXIDES ON PAPER AND ARCHIVAL
9 MATERIALS
10 Paper is the primary storage medium for permanent records ranging from personal
11 photographs to the Constitution of the United States. The National Research Council (1986)
12 noted that NO2 and other "acid gases" are expected to promote the failure of the cellulose
13 fibers that make up paper. They recommended that the storage condition standards suggested
14 by the National Institute of Standards and Technology be followed and that NOX levels in
* ' ^l
15 archives, libraries, and museums not exceed 5 jug/m .
16 Baer and Banks (1985) have pointed out a particular problem with NOX pollution that
17 libraries, museums, and archives face. In the nineteenth century, cellulose nitrate was
18 produced in large quantities as the first plastic and was used in a wide variety of products.
19 The common uses included photographic film, "acetate" recording disks, pre-vinyl imitation
20 leather, adhesives, and finishes. As cellulose nitrate ages it continuously emits NOX. If large
21 quantities of books with artificial leather bindings (or rebinding using pyroxylin-impregnated
22 cloth) or of early photographic film are stored, NOX indoor emissions, which can be
23 significant, may cause elevated concentrations unless the storage area is adequately vented.
24 In extreme cases of nitrate film storage in sealed vaults with no ventilation, the resulting gas
25 pressure "may be enough to force out masonry walls". If cellulose nitrate film is stored in
26 sealed containers, NOX concentrations can buildup to the point of causing an autocatalytic
27 reaction that can end in spontaneous combustion. Several collections of historic motion
28 picture films have been destroyed in fires resulting from this process.
29 Salmon et al. (1990) measured nitrogen species deposition during two seasons in five
30 museums in Los Angeles and measured outdoor concentrations of NOX species, as well.
31 They noted that previous studies that attributed the damage to NO2 may have actually been
August 1991
12-27 DRAFT - DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
seeing damage induced by "co-pollutant" species, such as HNO3. Concentrations of HNO3
within the museums were in the range of 1 to 40% of the outdoor concentrations. They
measured apparent HNO3 deposition velocities to vertical surfaces inside the museums, and
found values of approximately 0.18 to 2.37 cm/s. They suggested that the deposition of total
inorganic nitrate (gas-phase plus aerosol-phase) onto vertical surfaces is dominated by gas-
phase species (probably HNO3 vapor). A further study of HNO3 removal by air-handling
systems was conducted at one museum, and Salmon et al. (1990) found that approximately
40% of the HNO3 was removed by deposition within the ventilation system. It was
suggested that very low measured values of HNO3 within galleries may be misleading.
Deposition of HNO3 on surfaces within the museums, probably including the collection, was
rapid and potentially induced damage.
12.8 COSTS OF MATERIALS DAMAGE FROM NITROGEN OXIDES
Cost estimates for materials damage have been based on two distinct approaches. The
first technique, the "top-down approach," involves determining the dollar value of a material
produced each year and then estimating the percentage of that value that is lost each year
from pollutant-induced damage. The advantage of this approach is its ease of application.
However, it is not rigorous and is likely to contain significant errors. For example, using the
top-down approach it is not possible to determine the pollutant exposure levels of the
materials since mere is no way to determine the locations in which the materials are
deployed. All that can be done is to use gross averages for exposures with this technique.
The second technique is the "bottom-up approach," in which as much detail as possible
is gathered regarding the geographic distribution of materials, the spatially resolved pollutant
concentrations and other variables, and the costs of repairs and replacement. The bottom-up
approach is more rigorous and demanding in terms of data requirements, and may yield a
closer estimate of actual costs than the production approach. The accuracy of either approach
is unknown. The methodology of cost estimation for materials damage needs further research
and development.
August 1991
12-28 DRAFT - DO NOT QUOTE OR CITE
-------
1 The costs of some types of NOx-induced damage to textiles were estimated by the
2 National Research Council (1977). The following estimates, in 1977 dollars and based on
3 1977 production rates, were made.
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20..
21
22
23
24
25"
26
27
28
29
30
31
32
33
34
35
36
1. . $53 million incurred from dye fading on acetate fibers. This
includes costs for more expensive, fade-resistant dyes; inhibitors;
research; quality control; fade losses at the manufacturing and retail
level; and reduced product life at the consumer level as the result of
fading. ^ , •.. ,
2. $22 million incurred from dye fading on cotton fibers. This includes
estimates of cotton fabrics exposed in polluted areas, percentages of
dyes known to be susceptible to NOX fading, and yearly loss in use-
life.
3. $22 million incurred from dye fading on viscose rayon and rayon
blends with nylon, polyester, or acetate. This includes reduced wear-
life for sensitive dye shades. -...•-.
Estimates of the costs of other types of losses caused by adverse NOX impacts on
textiles and fibers are not available. Loss of strength and shortened use-life may be a
significant cost for fibers used for industrial purposes. According to the National Research
Council (1977), 18 to 20% of all fibers produced are used by industry for items such as
tarpaulins, cords, and rope. Loss of strength for fibers used for these'purposes shortens use-
life and may present a safety hazard.
Estimates of the costs of NOx-induced damage to plastics and elastomers are not
reported in the literature. The damages suffered through cross-linking and chain-scissioning
are loss of strength, increased cracking, and discoloration. As the use'of these compounds
for construction and automotive applications increases, the amount of exposure to NOX will
increase and the disbenefit costs of this exposure are expected to increase.
No overall estimates of the costs of NOx-mduced damage to metals and electronics are
available. For metallic corrosion in general, the" costs are large. The paint and coatings
industry, for example, produces a spectrum of products designed to prevent rust on steel and
these coatings would not be needed if corrosion were not a problem.
.August 1991
12-29 DRAFT - DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
Damage to paints, concrete, and stone produces potentially one of the largest economic
disbenefits of NOx-induced materials damage because the use of these materials is
widespread. In 1987, sales by the paints and coatings industry alone approached $10 billion.
The costs of infrastructure replacement because of concrete degradation can be seen as part of
the annual highway budgets. Damage to historic stone structures and statues is mostly a
cultural cost and is not readily calculated. Jirillo et al. (1987), however, have reported the
costs of preservation of Italian artistic properties as 69,697 thousand million lira, with a
sizable fraction of the damage attributed by the authors to acid deposition.
12.9 SUMMARY OF THE EFFECTS OF NITROGEN OXIDES ON
MATERIALS
Nitrogen oxides have been shown to cause or accelerate damage to manmade materials
exposed to the atmosphere. Strong evidence exists for the negative impact of NO on dyes
and fabrics. Many varieties of dyes are known to fade, become duller, or acquire a different
cast, and white fabrics may "yellow" from exposure to NOX. Fade-resistant dyes and
inhibitors have been developed, but are generally more costly to employ. Nitrogen oxides
also attack textile fibers and result in a loss of strength. Plastics and elastomers are subject to
NOX reactions that cause discoloration and changes in physical properties, including loss of
strength. The rate of NOx-induced deterioration to plastics and elastomers is generally
overshadowed, however, by O3-induced damage. Although NOX attacks metals, attack by
SO2 is more aggressive. Damage to metals from NOX can generally be discounted, except
perhaps in indoor exposures, where NOX may react synergistically with SO2. Also largely
indoors, NOX is deposited on electronic components and magnetic recording equipment and
may lead to failures in these systems. The influence of NOX on paints and stone has not been
clearly demonstrated. Many authors indicate that NOX plays a role in damaging these
materials, but most concede that SO2 and O3 are more directly damaging. Nitrogen oxides,
along with other "acid pollutants," attack the cellulose fibers in paper, leading to
discoloration and weakened structure.
The presence of NOX will shorten the use-life of susceptible, materials and generally the
rate of damage is proportional to the pollutant concentration. Adequate NOX damage
August 1991
12-30 ; DRAFT - DO NOT QUOTE OR CITE
-------
• 1, functions for a wide variety of materials are not available. Consequently, cost/benefit
2 analyses of permissible NOX levels vis-a-vis shortened use-life estimates could be misleading.
3 Cost estimates for NOx-specific damage at existing concentrations are available only for dye
4 fading ($97 million annually in 1977 dollars), and these estimates are very out of date.
5 The highest NOX levels are to be found indoors where unvented combustion systems
6 (e.g., gas stoves) are used and the widest variety of materials are routinely exposed.
7 Therefore, the principal effects of NOx-induced damage to materials are probably seen in the
8 indoor environment, but few data are available regarding materials deterioration indoors.
9 This is an area of needed research.
10 . . . .
August 1991
12-31 DRAFT - DO NOT QUOTE OR CITE
-------
REFERENCES
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44
45
46
47
48
49
50
51
52
53
Abbott, W. H. (1987) Corrosion of porous gold plating in field and laboratory environments. Plat. Surf. Finish.
74:72-75.
Amoroso, G. G.; Fassina, V. (1983) Stone decay and conservation: atmospheric pollution, cleaning,
consolidation and protection. Amsterdam, The Netherlands: Elsevier Science Publishers B.V. (Materials
science monographs 11).
Baedecker, P. A.; Edney, E. O.; Morgan, P. J.; Simpson, T..C.; Williams, R. S.; Hosker, R. P.;
Longmuir, D.; Mossotti, V. G.; McGee, E. S.; Reddy, M. M.; Spiker, E. C.; Weseley, M. J.;
Kishiyama, G.; Pavich, M. V.; Raimann, K. J.; Schmiermund, R. L.; Sciammarella, C. A.;
Youngdahl, C. A. (1990) Effects of acidic deposition on materials. Washington, DC: National.Acid
Precipitation Assessment Program; State of science and technology report no. 19.
Baer, N. S.; Banks, P. N. (1985) Indoor air pollution: effects on cultural and historic materials. Int. J. Museum,
Manage. Curatorship 4: 9-20.
Bassett, M.; Seinfeld, J. H. (1983) Atmospheric equilibrium model of sulfate and nitrate aerosols. Atmos.
Environ. 17: 2237-2252.
Baumgaertaer, M.; Remde, A.; Bock, E.; Conrad, R. (1990) Release of nitric oxide from building stones into
the atmosphere. Atmos. Environ. Part B 24: 87-92.
Beloin, N. J. (1972) Fading of dyed fabrics by air pollution. Text. Chem. Color. 4: 77-82.
Beloin, N. J. (1973) Fading of dyed fabrics exposed to air pollutants. Text. Chem. Color. 5: 128-133.
Cramer, S.; Carter, J. P.; Linstrom, P. J.; Flinn, D. R. (1988) Environmental effects in the atmospheric
corrosion of zinc. la: Dean, S. W.; Lee, T. S., eds. Degradation of metals in the atmosphere:
proceedings of a symposium on corrosion of metals; May 1986; Philadelphia, PA. Philadelphia, PA:
American Society for Testing and Materials; pp. 229-247. (ASTM special technical publication 965).
Edney, E. O.; Stiles, D. C.; Spence, J. W.; Haynie, F. H.; Wilson, W. E. (1986) A laboratory study to evaluate
the impact of NOX, SOX, and oxidants on atmospheric corrosion of galvanized steel. In: Baboian, R., ed.
Materials degradation caused by acid rain: developed from the 20th state-of-the-art symposium of the
American Chemical Society; June 1985; Arlington, VA. Washington, DC: American Chemical Society; ."
pp. 172-193. (ACS symposium series 318). . .
j
Edney, E. O.; Stiles, D. C.; Corse, E. W.; Wheeler, M. L.; Spence, J. W.; Haynie, F. H.; Wilson, W. E., Jr..
(1987) Field study to determine the impact of air pollutants on the corrosion of galvanized steel.,
Presented at: Corrosion/87; San Francisco, CA. NACE; paper no. 410.
Franey, J. P.; Graedel, T. E. (1985) Corrosive effects of mixtures of pollutants. J. Air Pollut. Control Assoc.
35: 644-648.
Freitag, W. O.; Mee, P.; Petersen, R. (1980) Glass/ferrite interactions and corrosion of gap glasses in recording.
heads. IEEE Trans. Magn. 16: 876-878.
Graedel, T. E.; McGill, R. (1986) Degradation of materials in the atmosphere. Environ. Sci. Technol.
20: 1093-1100.
Hawley, G. G. (1981) The condensed chemical dictionary. 10th ed. New York, NY:' Van Nostrand Reinhold
Company.
August 1991
12-32 DRAFT - DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44
45
46
47
48
49
50
51
52
53
Haynie, F. H. (1986) Environmental factors affecting corrosion of weathering steel. In: Baboian, R., ed.
'Materials degradation caused by acid rain: developed from the 20th state-of-the-art symposium of the
American Chemical Society; June 1985; Arlington, VA. Washington, DC: American Chemical Society;
pp. 163-171. (ACS symposium series 318).
Haynie, F. H.; Spence, J. W. (1984) Air pollution damage to exterior household paints. J. Air Pollut. Control
Assoc. 34: 941-944.
Haynie, F. H.; Upham, J. B. (1974) Correlation between corrosion behavior of steel and atmospheric pollution '
'data. In: Corrosion in natural environments. Philadelphia, PA: American Society for Testing and
Materials; pp. 33-51. (ASTM special technical publication 558).
'Haynie, F. H.; Spence, J. W'.; Upham, J. B. (1976) Effects of gaseous pollutants on materials--a chamber study,
Research Triangle Park, NC: U. S. Environmental Protection Agency, Environmental Sciences Research
Laboratory; EPA report no. EPA-600/3-76/015. Available from: NTIS, Springfield, VA; PB-251580.
Hempbill, J. E.; Norton, J. E.; Ofjord, O. A.; Stone, R. L. (1976) Colorfastness to light and atmospheric
contaminants. Text. Chem. Color. 8: 60-62.
Henriksen, J. F.; Rode, A. (1986) Corrosion rates of various metals in SO2/NO2 polluted atmospheres. In:
Proceedings of the 10th Scandinavian corrosion congress; June; Stockholm, Sweden. Stockholm, Sweden:
Swedish Corrosion Institute; pp. 39-42.
Hermance, H. W. (1966) Combatting the effects of smog on wire-spring relays. Bell Lab. Rec. (February):
48-52. '
Hermance, H. W.; Russell, C. A.; Bauer, E.' J.; Egan, T. F.; Wadlow, H. V. (1971) Relation of airborne nitrate
to telephone equipment damage. Environ. Sci. Technol. 5: 781-785.
Imperial Chemical Industries Limited. (1973) Nylomine dyes: effect of burnt gas fumes. Manchester, United
Kingdom: Imperial Chemical Industries Limited, Organics Division; publication no. D 1322, synthetic
fibre dyeing.
Jellinek, H. H. G. (1970) Chain scission of polymers by small concentrations (1 to 5 ppm) of sulfur dioxide and
nitrogen dioxide, respectively, in presence of air and near ultraviolet radiation. J. Air Pollut. Control
Assoc. 20: 672-674.
Jellinek, H. H. G.; Flajsman, F.; Kryman, F. J. (1969) Reaction of SO2 and NO2 with polymers. J. Appl.
Polym. Sci. 13: 107-116.
Jirillo, R.; Fonseca, G.; Ruggeri, P. (1987) Effects of acid rain on the environment. Some technical and
economic aspects of the Italian situation. In: Perry, R., ed. Aci'd rain. London, United Kingdom: Selper
Publishers; pp. 801-807.
Joerg, F.; Schmitt, D.; Ziegahn, K.-F. (1987) Materialschaden durch Lufrverunreinigungen. Sohderdr.
Landsburg/Lech: Economed.
Johansson, L.-G. (1986) A laboratory study of the influence of NO2 and SO2 on the atmospheric corrosion of
steel, copper, zinc and aluminium. Proc. Electrochem. Soc. 86-6: 267-279.
Johansson, L.-G.; Lindqvist, O.; Mangio, R. E. (1988) Corrosion of calcareous stones in humid air containing
SO2 and NO2. Durability Build. Mater. 5: 439-449.
August 1991
12-33 DRAFT - DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44
45
46
47
48
49
50
51
52
53
54
Krause, H. H.j Bohn, M. A.; Pfeil, A.; Schmitt, D.; Ziegahn, K.-F. (1989) Polymer failures induced by air
pollutants. In: Brasser, L. J.; Mulder, W. C., eds. Man and his ecosystem: proceedings of the 8th world
clean air congress, v. 2; September; The Hague, The Netherlands. Amsterdam, The Netherlands: Elsevier
Science Publishers B.V.; pp. 337-342.
Kucera, V. (1986) Influence of acid deposition on atmospheric corrosion of metals: a review. In: Baboian, R.,
ed. Materials degradation caused by acid rain: developed from the 20th state-of-the-art symposium of the
American Chemical Society; June 1985; Arlington, VA. Washington, DC: American Chemical Society;
pp. 104-118. (ACS symposium series 318). :
Lipfert, F. W. (1989) Dry deposition velocity as an indicator for SO2 damage to materials. JAPCA 39: 446-452.
Livingston, R. A. (1985) The role of nitrogen oxides in the deterioration of carbonate stone. In: Felix, G., ed.
Proceedings of the 5th international congress on deterioration and conservation of stone; Lausanne,
Switzerland. Lausanne, Switzerland: Presses Polytechniques Romandes; v. 1; pp. 509-516.:
Livingston, R. A.; Baer, N. S. (1983) Mechanisms of air pollution-induced damage to stone. In: Proceedings of
the Vlth world congress on air quality, v. 3, atmospheric pollution and materials; May; Paris, France.
Paris, France: International Union of Air Pollution Prevention Associations; pp. 33-40.
Loskutov, A. L; Enikeev, E. Kh.; Rozenfel'd, I. L.; Alekseev, V. N. (1982) Interaction of a nitrogen
dioxide-oxygen-water vapor mixture with oxidized surfaces of aluminum and iron. Prot. Met (Engl
Transl.) 18: 183-186.
Mansfeld, F. (1980) Regional air pollution study: effects of airborne sulfur pollutants on materials. Research •
Triangle Park, NC: U. S. Environmental Protection Agency, Environmental Sciences Research
Laboratory; EPA report no. EPA-600/4-80-007. Available from: NTIS, Springfield, VA; PB81-126351.
McKinney, N.; Hermance, H. W. (1967) Stress corrosion cracking rates of a nickel-brass alloy under applied
potential. In: Stress corrosion testing: a symposium presented at the sixty-ninth annual meeting of the
American Society for Testing and Materials; June-July 1966; Atlantic City, NJ. Philadelphia, PA:
American Society for Testing and Materials; pp. 274-291. (ASTM special technical publication no. 425).
McLendon, V.; Richardson, F. (1965) Oxides of nitrogen as a factor in color changes of used and laundered
cotton articles. Am. Dyest. Rep. 54: 15-21. , -
Miyazaki, T. (1984) Adsorption characteristics of NOx by several kinds of interior materials. In: Berglund, B.;
Lindvall, T.; Sundell, J., eds. Indoor air: proceedings of the 3rd international conference on indoor air
quality and climate, v. 4, chemical characterization and personal exposure; August; Stockholm, Sweden.
Stockholm, Sweden: Swedish Council for Building Research; pp. 103-110. Available from: NTIS
Springfield, VA; PB85-104214.
National Research Council. (1977) Nitrogen oxides. Washington, DC: National Academy of Sciences.
National Research Council. (1986) Preservation of historical records. Washington, DC: National Academy Press.
Rice, D. W.; Cappell, R. J.; Kinsolving, W.; Laskowski, J. J. (1980a) Indoor corrosion of metals. £'
J. Electrochem. Soc. 127: 891-901. ,«
Rice, D. W.; Phipps, P. B. P.; Tremoureux, R. (1980b) Atmospheric corrosion of nickel. J. Electrochem. Soc :
127: 563-568.
Rice, D. W.; Peterson, P.; Rigby, E. B.; Phipps, P. B. P.; Cappell, R. J.; Tremoureux, R. (1981) Atmospheric
corrosion of copper and silver. J. Electrochem. Soc. 128: 275-284.
August 1991
12-34 DRAFT - DO NOT QUOTE OR CITE
-------
1
2 ,
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44
45
46
47
48
49
50
51
52
53
54
Rowe, F. M.; Chamberlain, K. A. J. (1937) The "fading" of dyeings on cellulose acetate rayon: the action of
"burnt gas fumes" (oxides of nitrogen, etc. in the atmosphere) on cellulose acetate rayon dyes. J. Soc.
Dyers Colour. 53: 268-278.
Salmon, L. G.; Nazaroff, W. W.; Ligocki, M. P.; Jones, M. C.; Cass, G. R. (1990) Nitric acid concentrations
in southern California museums. Environ. Sci. Technol. 24: 1004-1013.
Salvin, V. S. (1964) Relation of atmospheric contaminants and ozone to lightfastaess. Am. Dyest. Rep.
53: 33-41.
Salvin, V. S. (1969) Testing atmospheric fading of dyed cotton and rayon. Am. Dyest. Rep. 58: 28-29.
Salvin, V. S.; Walker, R. A. (1959) Relation of dye structure to properties of disperse dyes:
part I—anthraquinone blues. Am. Dyest. Rep. 48: 35-43.
•N
Salvin, V. S.; Paist, W. D\; Myles, W. J. (1952) Advances in theoretical and practical studies of gas fading.
Am. Dyest. Rep. 14: 297-304.
Schubert, R. (1978) Interaction of NOX and Cu at various relative humidities. J. Electrochem. Soc.
125: 1114-1116.
Sherwood, S.; Davidson, C.; Dolske, D.; Gatz, D.; Hicks, B.; Hosker, R.; Langmuir, D.; Linzey, R.;
Lipfert, F.; Mossotti, V. G.; Schmiermund, R. L.; Spiker, E. C. (1990) Processes of deposition to
structures. Washington, DC: National Acid Precipitation Assessment Program; State of science and
technology report no. 20.
Spence, J. W.; Haynie, F.; Upham, J. B. (1975) Effects of gaseous pollutants on paints: a chamber study.
J. Paint Technol. 47: 57-63.
Spence, J. W.; Haynie, F. H.; Stiles, D. C.; Edney, E. O. (1988) A study of the effects of dry and wet
deposition on galvanized steel and weathering steel: a three-year field exposure. In: Dean, S. W.;
Lee, T. S., eds. Degradation of metals in the atmosphere: proceedings of the symposium on corrosion of
metals; May 1986; Philadelphia, PA. Philadelphia, PA: American Society for Testing and Materials;
pp. 339-353. (ASTM special technical publication 965).
Spicer, C. W.; Coutant, R. W.; Ward, G. F.; Joseph, D. W.; Gaynor, A. J.; Billick, I. H. (1987) Rates and
mechanisms of NO2 removal from indoor air by residential materials. In: Seifert, B.; Esdorn, H.;
Fischer, M.; Rueden, H.; Wegner, J., eds. Indoor air '87: proceedings of the 4th international
conference on indoor air quality and climate, v. 1, volatile organic compounds, combustion gases,
particles and fibres, microbiological agents; August; Berlin, Federal Republic of Germany. Berlin,
Federal Republic of Germany: Institute for Water, Soil and Air Hygiene; pp. 371-375.
Svedung, O. A.; Johansson, L.-G.; Vannerberg, N.-G. (1983) Corrosion of gold-coated contact materials
exposed .to humid atmospheres containing low concentrations of SO2 and NO2. IEEE Trans. Compon.
Hybrids Manuf. Technol. CHMT-6: 349-355.
Tombach, I. (1982) Measurement of local climatological and air pollution factors affecting stone decay. In:
Conservation of historic stone buildings and monuments: report of the Committee on Conservation of
Historic Stone Buildings and Monuments. Washington, DC: National Academy Press; pp. 197-210.
Vijayakumar, R.; Mansfeld, F.; Henry, R. (1989) Investigation of the effects of acid deposition on materials
[final report]. Camarillo, CA: Combustion Engineering, Environmental Monitoring 'and Services, Inc.;
contract nos. A4-110-32 and A5-137-32 for California Air Resources Board, Sacramento, CA.
August 1991
12-35 DRAFT - DO NOT QUOTE OR CITE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
Voytko, J. E.; Guilinger, T. R. (1988) Effectiveness of electrodeposited coatings in reducing atmospheric
corrosion of copper substrates. Proc. AESF Annu. Tech. Conf. 75: C-3, llpp.
Webster, R. P.; Kukacka, L. E. (1985) Effects of acid deposition on the properties of Portland cement concrete
state-of-knowledge. Research Triangle Park, NC: U. S. Environmental Protection Agency, Atmospheric
Sciences Research Laboratory; EPA report no. EPA-600/3-85/011. Available from: NTIS, Springfield,
VA; PB85-171452.
Whitbeck, M. R.; Jones, D. A. (1987) Effects of temperature, humidity, nitrogen dioxide and nitric acid gases
on carbon-steel, galvanized, and painted steel. Reno, NV: University of Nevada, Water Resource Center;
report no. ARB-R-87/310. Available from: NTIS; Springfield, VA; PB88-130661.
Yocom, J. E.; Baer, N. S. (1983) Effects on materials: the acid deposition phenomenon and its effects.
Washington, DC: U. S. Environmental Protection Agency, Office of Research and Development; EPA
report no. EPA-600/8-83-016B.
Zeronian, S. H.; Alger, K. W.; Omaye, S. T. (1971) Reaction of fabrics made from synthetic fibers to air
contaminated with nitrogen dioxide, ozone, or sulfur dioxide. In: Englund, H. M.; Beery, W. T., eds.
Proceedings of the second international clean air congress; December 1970; Washington, DC.
New York, NY: Academic Press; pp. 468-476.
*U.S. GOVERNMENT PRINTING OFFICE: 1W2-64S-003>40671
August 1991
12-36 DRAFT - DO NOT QUOTE OR CITE
-------
-------
-------
-------
m
"U
Q) O
CD
O •<
p _
So o 5
«b " '-
;:• n>
I a
to
»
> m i
(Q D
(I) <
II
3
CD
T)
s
(D
O
o'
3
Q. =•
.
c±. o
0 =
-------