RESEARCH
A Conceptual Model for Assessing Ecological Risk
to Water Quality Function of Bottomland
Hardwood Forests
RICHARD LOWRANCE
USDA-ARS
Southeast Watershed Research Laboratory
Tifton, Georgia 31793, USA
GEORGE VELLIDIS
Biological & Agricultural Engineering Department
University of Georgia
Tifton, Georgia 31793, USA
ABSTRACT / Ecological risk assessment provides a
methodology for evaluating the threats to ecosystem
function associated with environmental perturbations or
stressors. This report documents the development of a
conceptual model for assessing the ecological risk to the
water quality function (WQF) of bottomland hardwood
riparian ecosystems (BHRE) in the Tifton-Vidalia upland
(TVU) ecoregion of Georgia. Previous research has
demonstrated that mature BHRE are essential to
maintaining water quality in this portion of the coastal
plain. The WQF of these ecosystems is considered an
assessment endpoint--an ecosystem function or set of
Society generally understands the importance of
wetland ecosystems as part of the human life-support
system--the environment, organisms, processes, and
resources interacting to provide t h e physiological necessities of life (Odum 1989). This understanding is
embodied in a set of rules, regulations, and voluntary
common usage, which, taken together, provides some
protection for wetland ecosystems. Although afforded some protection, the life-support functions of
wetlands are at risk due to a variety o f factors including logging, conversion to agricultural lands and
overloading of pollutants. Given that society values
the life-support role of wetland ecosystems, the
KEY WORDS: Ecological risk assessment; Conceptual model; Bottomland hardwood forests; Water quality function;
Stressors; Assessment endpoint; Measurementend
point; Georgia
functions that society chooses to value as evidenced by
laws, regulations, or common usage. Stressors operate on
ecosystems at risk through an exposure scenario to
produce ecological effects that are linked to loss of the
desired function or assessment end point. The WQF of
BHRE is at risk because of the ecological and
environmental quality effects of a suite of chemical,
physical, and biological stressors. The stressors are
related to nonpoint source pollution from adjacent land
uses, especially agriculture; the conversion of BHRE to
other land uses; and the encroachment of domestic
animals into BHRE. Potential chemical, physical, and
biological stressors to BHRE are identified, and the
methodology for evaluating appropriate exposure
scenarios is discussed. Field-scale and watershed-scale
measurement end points of most use in assessing the
effects of stressors on the WQF are identified and
discussed. The product of this study is a conceptual
model of how risks to the WQF of BHRE are produced and
how the risk and associated uncertainties can be
quantified.
threats to life support need to be quantitatively assessed and the results o f the assessment need to be
used to develop risk management strategies to assure
maintenance o f these functions. In this paper, we develop a conceptual model for how the risk to one
particular set o f wetland ecosystems, the bottomland
hardwood riparian ecosystems (BHRE), in an important agricultural region o f the southeastern United
States can be quantitatively assessed. T o develop this
conceptual model, we draw on literature sources to
determine what stressors will potentially affect the
ability o f these wetland ecosystems to control nonpoint source pollution and how the impacts o f these
stressors can be measured and integrated to produce
the framework for a quantitative risk assessment.
Ecological Risk Assessment
*Author to whom correspondenceshould be addressed.
A Framework for Ecological Risk Assessment has been
developed by the United States Environmental Pro-
Environmental Management Vol. 19, No. 2, pp. 239-258
9 1995 Springer-Verlag New York Inc.
240
R. Lowrance and G. Vellidis
Table 1.
Definitions of ecological risk assessment terms as used in the conceptual model
Key terms
Definitions
Assessment end point
The ecosystem value or function that is to be protected and that is thought to
be at risk.
Measurable response to an actual stressor that is related to the valued
characteristic(s) chosen as the assessment end point(s).
Any physical, chemical, or biological entity that can cause an adverse ecological
effect.
A description of how the potential stressors contact the ecosystems at risk.
Stressors that, based on the exposure scenario, are likely to contact the
ecosystem(s) providing the desired or valued function.
The adverse effects elicited by a stressor.
Ecosystems that provide the func(ion embodied in the assessment end point
and that are likely to contact stressors.
Hypotheses concerning the impacts of actual stressors on ecosystems at risk and
how the resuhing ecological effects create a risk to the assessment end point.
Measurement end point
Potential stressor
Exposure scenario
Actual stressors
Ecological effect
Ecosystems at risk
Conceptual model
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QUANTITATIVE RISK ANALYSIS
RISK C H A R A C T E R I Z A T I O N
Fig. 1. The three phases of the BHRE ecological risk assessment process with the problem formulation phase presented in detail (adapted from Hunsaker and others 1990).
tection A g e n c y (USEPA) to h e l p clarify the steps
t a k e n in ecological risk a s s e s s m e n t a n d to p r o v i d e a set
o f c o m m o n t e r m s for t h e discussion o f risk a s s e s s m e n t
[Risk A s s e s s m e n t F o r u m R A F 1992]. I n t e r p r e t a t i o n s
o f these c o m m o n t e r m s , b a s e d on b o t h tile f r a m e w o r k
a n d o u r use in this article, a r e given in T a b l e 1. Ecological risk a s s e s s m e n t is d e f i n e d as a process to evaluate the p r o b a b i l i t y that " a d v e r s e ecological effects m a y
o c c u r o r a r e o c c u r r i n g as a result o f e x p o s u r e to o n e
o r m o r e stressors" ( R A F 1992). In the f r a m e w o r k , risk
assessment is a t h r e e - s t e p process: (l) p r o b l e m f o r m u lation; (2) q u a n t i t a t i v e risk analysis; a n d (3) risk d e scription a n d i n t e g r a t i o n . F i g u r e 1 shows the g e n e r a l
process o f f o r m u l a t i n g t h e p r o b l e m . T h e o u t c o m e o f
t h e p r o b l e m f o r m u l a t i o n p r o c e s s is a specific c o n c e p tual m o d e l that can b e u s e d to assess the risks to p a r ticular ecosystem functions.
T h e first s t e p in ecological risk a s s e s s m e n t is to
d e t e r m i n e an "assessment e n d point." A s s e s s m e n t
e n d p o i n t s a r e explicit e x p r e s s i o n s o f the actual ecosystem value o r l i f e - s u p p o r t f u n c t i o n that is to be p r o tected. S u t e r (1990) lists a n u m b e r o f characteristics o f
g o o d a s s e s s m e n t e n d points; t h e most i m p o r t a n t o f
these is social relevance. O n c e scientific k n o w l e d g e
a b o u t an ecosystem f u n c t i o n is m a d e available to society in a usable f o r m , if society d o e s n o t e x p r e s s interest in the loss o f t h e ecosystem f u n c t i o n t h r o u g h laws,
r e g u l a t i o n s , o r o t h e r m e a s u r e s , the p a r t i c u l a r ecosystem f u n c t i o n is not a g o o d a s s e s s m e n t e n d point. As
S u r e r (1990) points o u t *'we c a n n o t assess effects on
e v e r y t h i n g . " T h e s e c o n d m a j o r c h a r a c t e r i s t i c is that
the a s s e s s m e n t e n d p o i n t m u s t be susceptible to the
measurable and, one would hope, predictable impact
of" stressors. In o t h e r words, t h e a s s e s s m e n t e n d p o i n t
m u s t be at risk b e c a u s e o f i d e n t i f i a b l e ecological effects f r o m k n o w a b l e stressors.
T h e selection o f the a s s e s s m e n t e n d p o i n t is necessarily a c c o m p a n i e d by selection o f spatial a n d t e m p o ral b o u n d a r i e s f o r the risk a s s e s s m e n t ( F i g u r e 1)
( H u n s a k e r a n d o t h e r s 1990). Spatial b o u n d a r i e s a r e
g e n e r a l l y set by s o m e u n d e r s t a n d i n g o f t h e geog r a p h i c e x t e n t o f w h e r e an ecosystem f u n c t i o n is imp o r t a n t . T o simplify the risk assessment, the spatial
b o u n d a r i e s a r e likely to be a r e g i o n w h e r e t h e r e a r e
relatively h o m o g e n o u s e n v i r o n m e n t a l c o n d i t i o n s cont r o l l i n g the ecosystem f u n c t i o n s o f interest. T e m p o r a l
b o u n d a r i e s a r e m o r e difficult to d e f i n e , especially because o f the ability o f at risk ecosystems to e i t h e r recover o r d e g r a d e o v e r time. T e m p o r a l b o u n d a r i e s a r e
m o r e likely t h a n spatial b o u n d a r i e s to be set arbitrarily b a s e d on logistical a n d f u n d i n g constraints.
Model to Assess Ecological Risk
at Risk
j
Ecological
Effects
Fig, 2. Stressors operate on ecosystems at risk through an
exposure scenario to produce ecological effects.
Potential stressors fall into three general categories: physical, chemical, and biological (RAF 1992).
Chemical stressors include organic and inorganic substances intentionally or inadvertently introduced into
the environment. Physical stressors include extremes
o f natural conditions, such as droughts or floods, as
well as physical destruction or alteration (RAF 1992).
Biological stessors include organisms introduced into
the affected ecosystems, which alter ecosystem structure, function, or processes. I f a given stressor is unlikely to operate on an ecosystem that has an ecosystem function at risk, a potential stressor does not
become an actual stressor. Conceptually, stressors operate on ecosystems potentially at risk to produce ecological effects. T h e stressors operate through an exposure scenario that is based on existing knowledge or
hypotheses about the intersection of ecosystems potentially at risk with the stressors (Figure 2). T h e ecological effects are measured through a series of measurement end points.
Measurement end points are measurable responses
to as tressor that are related to the valued characteristics chosen as the assessment end points (Surer 1990,
RAF 1992). Generally, the assessment end point cannot be m e a s u r e d directly. Measurement end points
are selected to be of the most use in quantifying the
effects of stressors on the assessment end point. This
is the most important characteristic of any measurement end p o i n t - - t h a t it provide objective information
about the assessment end point. Measurement end
points can actually correspond to the assessment end
point, but usually the assessment end point will be a
broader attribute of ecosystem function and/or structure than can not be measured through one parameter. T h e r e f o r e , values of the m e a s u r e m e n t end point
must correlate with or help predict values of the assessment end point (Suter 1990). Measurement end
points based on either biological structure or function
are likely to provide the most information about the
assessment end point, but they may also vary most in
both time and space and in response to management.
One possible paradox of choosing a good measurement end point is pointed out by some of the attributes selected by Suter (1990). T h r e e characteris-
241
tics o f good m e a s u r e m e n t end points are to have low
natural variability, to have b r o a d applicability, and to
be diagnostic of the stressor. Many m e a s u r e m e n t s o f
processes or dynamic chemical pools would be
broadly applicable and would be diagnostic o f stressors, but would have high natural variability in response to highly variable environmental conditions.
One example of this type of m e a s u r e m e n t end point
would be denitrification in soils (reduction o f nitrate
to nitrogen gas by soil bacteria). D e p e n d i n g on environmental conditions or type of chemical stressor,
denitrification rate or potential could be a readily
measurable indicator of either nitrate pollution, organic matter enrichment, or both. Denitrifying organisms are nearly ubiquitous in soils and subsoil, yet
denitrification rate is highly variable in time and
space, largely due to the influence of carbon-rich
anaerobic microsites in soils (Parkin 1987).
Conceptual Model
Functions of Bottomland Hardwood
Riparian Ecosystems
T h e potential ecological impacts on B H R E are important to society because of the loss of ecosystem
functions known to be associated with bottomland
hardwoods or riparian wetland ecosystems. Preston
and Bedford (1988) have summarized these functions
as (1) hydrologic function; (2) water quality function;
and (3) life-support function. T h e i r description o f the
water quality function (WQF) is "the capacity of wetlands to remove or transform excess nutrients, organic compounds, trace metals, sediment, and refractory chemicals from water as it moves downstream."
T h e concept of the WQF used in this r e p o r t will include not only the ability of B H R E to remove pollutants but also includes their ability to control the physical and chemical conditions o f the adjacent aquatic
ecosystem.
Bottomland hardwood riparian ecosystems in the
eastern United States compose a set o f wetland ecosystems that have been shown to be important sinks for
nonpoint source pollution in studies conducted in a
n u m b e r of states including Rhode Island, Maryland,
North Carolina, and Georgia (Groffman and others
1992, Jacobs and Gilliam 1985, Lowrance and others
1984b, Peterjohn and Correll 1984). By functioning
as pollutant sinks, B H RE provide'direct i m p r o v e m e n t
of water quality by keeping pollutants derived f r o m
adjacent land uses out of streams or by removing polutants that entered upstream. This pollution control
function is due to a n u m b e r of processes including
uptake of nutrients by vegetation, loss o f nitrogen to
the a t m o s p h e r e via denitrification, deposition of sedi-
242
R. Lowrance and G. Vellidis
ment and sediment-borne pollutants, sequestering of
pollutants in high organic matter soils, and retention
of nutrients due to groundwater storage in alluvial
aquifers and surface water storage in swamps (Lowrance 1991). In addition to the removal of pollutants,
BHRE influence stream water quality through direct
inputs of detritus to streams; shading of stream channels, which controls both light availability and stream
temperature; and stabilization of stream banks and
channels, reducing channel-derived sediments (Karr
and Schlosser 1978, Welsch 1991).
T h e two distinct aspects of the WQF have different
characteristics relative to other land uses in the watershed. T h e function as a pollutant sink is primarily
situational, derived from the ability to process pollutants from adjacent or upstream sources. If pollutant
sources do not exist (a rare occurrence in managed
landscapes), then this component of the WQF could
be unimportant. T h e second component of the WQF,
control of the chemical and physical environment of
the stream, is more of an inherent ecosystem function.
Regardless of adjacent land uses or sources of pollutants, the BHRE controls these properties of the
stream. This distinction between situational and inherent functions is an important consideration in assessing risk to the WQF. T h e function as a pollutant
sink may be both enhanced and more at risk as the
loadings from upstream or upslope are increased.
T h e function as regulator of stream conditions may
be of more importance if upstream areas are degraded but is not as much at risk to increased pollutant loadings, unless the loadings cause death or replacement of the streamside vegetation.
After selection o f the assessment end point, the
next step in the risk assessment is to determine spatial
and temporal boundaries for the assessment. T h e
term "bottomland hardwood forests" has been used
historically to describe the floodplain forests found
throughout the southeastern United States (Taylor
and others 1990). T h e vegetation, hydrology, soils,
and functional attributes of these ecosystems vary
widely with a common thread of hydrophytic trees,
hydric soils, and water regimes strongly influenced by
river or stream hydrology. Preston and Bedford
(1988) recommend using the "magnitude of exchanges among component wetlands in a watershed
or ecoregion as the basis for defining geographic
boundaries. A different and perhaps more relevant
spatial boundary for assessment of the WQF can be
drawn on the basis of common hydrology, geology,
and land uses. Within a relatively homogeneous area,
one can evaluate the risk to the WQF on individual
watersheds and on larger basins downstream, if ap-
propriate. T h e spatial boundary must include a lower
bound as well. T h e lower boundary would most appropriately be the watershed (Preston and Bedford
1988), except that watersheds are not consistently
used as the basis for regulatory decisions concerning
impacts of stressors on wetland function. These decisions are generally made at the site scale and the stressors operate at the site scale, necessitating assessment
of risk at the site scale. For stressors resulting from
agricultural management o f adjacent areas or wetlands themselves, there would be some rationale for
making the assessment at the farm scale.
For risk assessment of the WQF of bottomland
hardwood forests, any region where there is a reasonable understanding of the water quality role played by
BHRE and in which a quantifiable conceptual model
could be developed would provide suitable spatial
boundaries. T h e Tifton-Vidalia Upland (TVU), described in detail below, is one such region and will
provide the spatial boundaries for this conceptual
model. In the coastal plain of Georgia, except for the
large alluvial streams that drain the Piedmont
(Oconee, Ocmulgee, and Savannah rivers, especially)
almost all coastal plain rivers fall into the category o f
hlackwater river and swamp system. For purposes of
this report, we will include only the portions o f the
rivers and streams originating in the T V U and contained in the ecoregion. This includes the entire
length of the Little, Little Ocmulgee, and Ohoopee
rivers and the u p p e r reaches of the Alapaha, Canoochee, Satilla, and Withlacoochee rivers.
Temporal boundaries are harder to set and are
likely to be more arbitrary. One guideline for temporal boundaries for risk assessment of the WQF is provided by both short-term and longer-term hydrologic
dynamics. Watersheds in the T V U have periods of
high r u n o f f on both annual and longer cycles. Highest r u n o f f each year generally takes place in winter
months when soil water storage is full and large rainfall events occur. In addition, there are longer-term
extremes of r u n o f f and pollutant transport associated
with wet and dry years. T h e temporal scale for risk
assessment should be long enough to encompass these
hydrologic extremes. Other possible guides for the
temporal boundaries include the timing of recovery
of the ecosystems at risk. Using this to set temporal
boundaries would be confounded by the fact that recovery from different stressors would take place at
vastly different rates. For instance, recovery of the
WQF from timber harvest without hydrologic modification probably takes place in a few years. Recovery of
WQF from hydrologic modifications associated with
conversion to agricultural land might take decades or
Model to Assess Ecological Risk
longer. T h e timing of disturbances (especially wet season vs dry season) will also affect the WQF.
Defining spatial and temporal boundaries only
partially defines the spatial and temporal reference o f
the risk assessment. Stressors can impact the WQF at
either the site or watershed scale and have an impact
on downstream water quality at either scale. Measurements o f impact o f impairment o f downstream water
quality due to loss o f the WQF may be more meaningful at the outlets o f larger watersheds, but these measurements integrate a larger n u m b e r o f factors than
just the WQF of BHRE. T h r o u g h o u t the development of the conceptual model, we will assume that
therisk assessment would be relevant to at least two
scales: site and watershed. By site scale, we are primarily concerned with the interrelationships among
specific land use practices and the risk to the WQF
and off-site water quality due to a site scale integration
o f stressors. At the watershed scale, we are primarily
concerned with integrating the cumulative effects of
different site scale stressors on the downstream water
quality from a larger area.
Tifton-Vidalia Upland Ecoregion
T h e Tifton-Vidalia Upland is a physiographic subprovince o f the Gulf-Atlantic Coastal Plain, which has
relatively homogeneous geology, soils, parent materials, land use, agricultural management, and economic
and social patterns. T h e ecologic, economic, and social cohesiveness of the T V U makes it possible to consider the area an ecoregion--"a geographical province with a marked ecological and often cultural
unity" (Berg 1981). T h e TVU includes all or most o f
28 counties andparts o f 16 others in Georgia, an area
approximately 52,000 km 2. T h e T V U is drained both
by rivers that originate in the ecoregion and by major
rivers that originate in the Georgia and South Carolina Piedmont and cut through the T V U on their way
to the Atlantic Ocean (Figure 3).
T h e climate o f the T V U is humid subtropical
(Strahler 1975) providing abundant rainfall and a
long growing season. Average monthly temperatures
at Tifton, Georgia, range from 11~ in January to
27~ in July and August with a mean annual temperature o f 19.2~ and mean annual rainfall of 120.3 cm
(Batten 1980). Topographically, the T V U is an area
o f floodplains, river terraces, and gently sloping uplands. Bottomlands are nearly level and most valley
flanks are less than 5% slope although some slopes o f
5%-15% exist.
T h e T V U is underlain by the u p p e r part o f the
Hawthorne Formation, which is composed o f Miocene age sediments (Asmussen and Ritchie 1969).
243
T h e Hawthorne Formation forms an effective
aquiclude since it transmits water at a much lower rate
than the porous medium above it. T h e aquiclude
causes most infiltrated precipitation to halt its downward movement and move laterally to the stream
channels. T h e generalized hydrology and landscape
of the T V U is illustrated in Figure 4, taken from a
false color infrared image o f Georgia. T h e dense dendritic network o f stream channels is bordered by riparian forest wetlands (dark areas) with the upland
areas devoted to mostly agricultural uses (bare
ground in white). Surface waters are primarily used
for irrigation, fisheries, and recreation. Shallow
groundwater is used to recharge farm ponds and for
limited irrigation and domestic water supply. Deeper
groundwater from the Principal Artesian Aquifer is
used for municipal and rural domestic water supply,
irrigation, and industries in this portion o f the coastal
plain. Because infiltrated rainfall cannot move effectively beneath the Hawthorne formation, recharge to
the Principal Artesian Aquifer below the Hawthorne
is minimal within the TVU.
Soils of the TVU are formed primarily from the
Hawthorne Formation with minor areas formed from
eolian sands. Most of the upland soils are classified as
fine-loamy (or loamy), siliceous, thermic Plinthic Kandiudults. T h e bottomland soils are primarily loamy,
siliceous, thermic Arenic Plinthis Paleaquults with
some Fiuvaquents and Psammaquents (Calhoun
1983). Most o f the upland soils contain plinthite, a
"sesquioxide-rich, humus-poor, highly weathered
mixture o f clay with quartz and other diluents" (Calhoun 1983), which forms an aquitard and can cause
perched water tables.
Vegetation o f the T V U has changed drastically
from presettlement times due to logging, agriculture,
and silviculture. T h e entire region was once called
"wiregrass country" because o f the longleaf pine/
perennial wiregrass community (Harper 1906). Little
of the original upland vegetation exists within the
TVU. Most upland areas have been converted to either agricultural uses or pine plantations. Many
poorly drained areas o f the ecoregion are still occupied by vetetation similar to the original presettlement plant communities. These wetland areas will be
described in detail in the section on Ecosystems Potentially at Risk (below).
Present land uses in the T V I ) are dominated by
agriculture and forestry. T h e land in farms accounts
for about 41% of the total area o f the T V U countries
(1987 Census o f Agriculture 1989) (Table 2). Total
cropland in the T V U counties is about 24% o f the
land area. T h e remaining land in farms is primarily in
244
I
R, Lowrance and G. Vellidis
LEGEND
Tifton-Vid<~fio Uplond
SCAI-E
0
\
25
miles
50
OCO
O/4
Fig. 3. The stream network and the
portions of the rivers originating and
contained in the TVU ecoregion.
Fig. 4. Satelite image of the TVU
showing the dense dendritic network of stream channels bordered
by riparian forests. Forests appear
dark gray or black and exposed/
cultivated earth appears light gray
or white.
w o o d l a n d (13% o f T V U ) o r p a s t u r e l a n d (2.5% of
T V U ) or other uses (6.2% of TVU). The 59% of the
T V U c o u n t i e s that is not in (~arms is p r i m a r i l y pri-
vately o w n e d forest l a n d with a b o u t 5% o f the T V U
counties in u r b a n , s u b u r b a n , r u r a l h o u s i n g , o r transp o r t a t i o n uses.
Table 2.
Model to Assess Ecological Risk
245
Percent
of 'FVU
land area
Percent
of TVU
farmland
Agricultural land use in the Tifton-Vidalia Upland ~
Percenl
Land-use category
Total land area (ha)
Forest land (ha)
Land drained, 1978 ~ (ha)
Land in farms, 1978 (ha)
Land in farms (ha)
Farms (number)
Farm avg size (ha)
Total cropland (ha)
Harvested cropland (ha)
Cropland-pasture/grazing (ha)
Other cropland (ha)
In cover, not harvested (ha)
All crops failed (ha)
Summer fallow (ha)
Idle (ha)
Total woodland (ha)
Woodland pastured (ha)
Woodland not pastured (ha)
Other land (ha)
Pastureland and rangeland (ha)
House lots, ponds, etc. (ha)
Pasturelands, all types (ha)
Diverted under commodity programs (ha)
In conservation reserve (ha)
Georgia
TVU
of
Georgia b
15,1/42,796
9,771,958
619,830
5,563,759
4,350,088
43,552
100
2,340,215
1,335,331
463,551
541,333
71,428
36,415
38,481
395,010
1,435,536
358,384
1,077,153
574,337
358,180
216,157
1,18(1,114
164,320
39,327
5,061,269
2,914,894
173,429
2,607,853
2,054,940
15,722
134
1,203,896
767,674
124,552
311,671
3 ! ,820
17,219
21,810
238,458
6411,672
120,769
5116,184
224,960
113,61)5
96,898
358,919
99,635
24,682
33.6
29.8
28.1/
46.9
47.2
36.1
133.7
51.4
57.5
26.9
57.6
44.5
47.3
56.7
60.4
44.6
33.7
47.0
39.2
31.7
44.8
30.4
60.6
62.8
100,0
57.6
3.4
51.5
4/).6
126.9 a
100.0
23.8
15.2
2.5
6.2
0.6
0.3
0.4
4.7
12.7
2.4
58.6
37.4
6.1
15.2
1.5
11.8
1.1
11.6
31.2
5.9
10.0
24,6
4.4
2.2
1.9
7.1
2.0
(1.5
10.9
5.5
4.7
17.5
4.8
1.2
q)ata are Jot 1987 unless otherwise slated. Data compiled from the 1987 Census of Agriculture (1989), the 1991 Georgia (:ounty Guide
(Bachtel and Boatright 1992), and the 1978 Census of Agriculture ( 1981).
~'Proportiou of Georgia included in the Tifton-Vidalia Upland tot each land use calegory.
~Most recent data available are from 1978.
dWith respect to land in farms in 1987.
T h e population o f the T V U grew by about 9% in
1980-1990. T h e growth was uneven, with the smallest
county shrinking by 9% f r o m its 1980 population and
the largest c o u n t y g r o w i n g by over 20%. In 1980,
41.4% o f the population was urban. T h e rural nonfarm population was the majority (51.1%) o f the population. Despite the i m p o r t a n c e o f agriculture to the
region, tile farm population was only 7.5% o f the
total. L o w r a n c e and Vellidis (1993) p r o v i d e d a m o r e
detailed description o f the T V U ecoregion.
Potential Stressors
B H R E are subject to ecological and e n v i r o n m e n t a l
d a m a g e d u e to a wide r a n g e o f h u m a n d e v e l o p m e n t
(Roelle and others 19901. Agricultural, suburban, and
industrial d e v e l o p m e n t pose particular problems for
b o t t o m l a n d hardwoods. Ecological impacts can be as
obvious as conversion o f b o t t o m l a n d h a r d w o o d s to
row crops t h r o u g h vegetation removal, drainage, and
soils modification, or they can be m o r e subtle impacts
such as n u t r i e n t enrichment. C o n c u r r e n t with the loss
o f these riparian ecosystems, there has been g r o w i n g
scientific evidence o f tile i m p o r t a n c e o f the Water
Quality Function o f b o t t o m l a n d h a r d w o o d s and riparian ecosystems in general ( G r o f f m a n and others
1992, Jacobs and Gilliam 1985, L o w r a n c e and others
1984b, Peterjohn and Correll 1984, S i m m o n s and
others 1992). T h e s e functions are especially a p p a r e n t
in areas d o m i n a t e d by land uses that i n t r o d u c e pollutants into the e n v i r o n m e n t . T w o areas where these
functional values are quite a p p a r e n t is in agricultural
or s u b u r b a n watersheds. Both agricultural and suburban d e v e l o p m e n t introduce stressors that have the
potential to d e g r a d e the functional value o f bottomland h a r d w o o d and riparian ecosystems. Potential
stressors are listed in Table 3.
Chemical stressors. T h e chemical stressors o f importance in tile T V U include plant nutrients (primarily
nitrogen and p h o s p h o r u s , including animal waste)
and pesticides (primarily insecticides, nematicides,
fungicides, and herbicides). T h e n o n p o i n t sources o f
these stressors in the T V U are primarily agriculture
246
R. Lowrance and G. Vellidis
Table 3. Potential stressors of BHRE in the
Tifton-Vidalia Upland
Chemical Stressors
Plant Nutrients
From nonpoint sources
Agriculture
Waterborne transport from proper
application
Direct release from improper application
Urban and suburban development
From point sources
Sewage treatment plants
Animal confinement facilities
Pesticides
From nonpoint sources
Agriculture
Waterborne transport from proper
application
Direct release from improper application
Spray drift
Urban and suburban development
From point sources
Formulation plants
Physical Stressors
Conversion to other vegetation
Drainage and soil modifications
For agricultural uses
For nonagricultural uses
Sedimentation
Biological Stressors
Livestock
Loss/damage of native vegetation
Direct introduction of waste products
Compaction
and urban/suburban development. The point sources
are primarily sewage treatment plants, animal confinement operations, and agrichemical formulation
plants.
Movement of agrichemicais into BHRE is primarily due to three processes: (1) waterborne transport
inboth surface runoff and groundwater flow; (2) direct release into BHRE by drift from proper application of sprays, powders, or granules; and (3) direct
release into BHRE due to improper application procedures. Application of nutrients and pesticides on
agricultural [ands does not necessarily mean that
there will be movement into BHRE, although transport of nutrients and pesticides beyond the edge of
fields does generally occur. Numerous models exist
for the prediction of loading rates of nonpoint source
pollution at the edge of fields. Given the landscape
structure common in the TVU, these edge-of-field
loadings can be considered as inputs to BHRE. Prediction of the potential loadings from nonpoint
source pollution will be discussed under Exposure
Scenarios, below.
Pesticide spray released from aircraft or ground
sprayers may, under various meteorologic conditions,
drift into areas of sensitive plant and animal species
(Himel and others 1990). For the purposes of this
report, drift is defined as those airborne components
that move downwind and outside the defined spray
area. Since almost all pesticides are volatile, drift consists of particulate pesticide and pesticide vapor.
Spray drift may result from aerial or ground application. Problems of drift are exacerbated by the fact that
only spray droplets <100--150 p.m deliver pesticides
into foliage environments. Larger droplets fall to peripheral foliage and the ground, decreasing spray efficacy and introducing active ingredient to nontarget
areas. The small droplets, however, are much more
prone to drift and evaporation.
Numerous studies conducted to quantify spray
drift (Himel and othes 1965, Barry and Ekbald 1978,
Uk 1977, Wiesner 1984) have determined that drift is
a probability problem driven by wind speed and direction, droplet density and size distribution, evaporation rate, volume and amount of pesticide applied,
and atmospheric turbulence, thus making it extremely difficult to predict inadvertent pesticide inputs to uontarget ecosystems. However, the potential
of nontarget BHRE receiving spray drift appears to
be large because target agricultural areas are commingled and contiguous to BHRE throughout the TVU
landscape.
The cause-and-effect relationship between pesticide use for agricultural production at a locale and
observation of agricultural chemical contamination of
the surrounding environment can not be readily established in many cases (Cheng 1990). However, reported incidents of adverse environmental or health
effects resulting from agricultural chemical contamination can often be traced to improper application or
inappropriate practices. Lack of knowledge or a disregard for the sensitivity of the environment is the root
of many agricultural chemical contamination problems (Cheng 1990). Injudicious application of highly
water-soluble pesticides to irrigated crops in sandy
soils, for example, could result in high concentrations
of pesticides in shallow groundwater underlying the
field. In the TVU landscape, this contaminated water
will likely pass through a BHRE before feeding a first,
second, or third-order stream.
Although the TVU is primarily a rural agricultural
and forestry region, urban/suburban development
and associated nonpoint and point sources of pollution have increased over the past decades. Data on
Model to Assess Ecological Risk
nonpoint source pollution loads from lawns, gardens,
golf courses, and other urban/suburban landscapes is
lacking. Some studies have indicated that there are
higher per hectare loadings o f nutrients and some
pesticides on gold courses than on typical agricultural
lands (Cisar and others 1991). Within the entire TVU,
the total loadings and potential stressor load from this
course would be considerably less than agricultural
loadings.
Point source pollution problems from sewage
treatment plants have been documented in at least
one area adjacent to the TVU. Pollutant loads from
the sewage treatment plant in Cordele, Georgia, are
known to have caused significant eutrophication of an
arm o f Lake Blackshear, a reservoir formed by damming o f the Flint River. T h e tributary receiving effluent from the sewage treatment plant is Gum Creek,
which originates in the TVU. Other documented
point source problems in the T V U that are potential
stressors for BHRE are releases o f pesticides from
formulation plants along tributary streams of the A1apaha and Little Rivers.
Physical stressors. Physical stressors include conversion o f BHRE to other vegetation, drainage and soil
modifications for agricultural and nonagricultural
uses, and sedimentation due to upland or upstream
sources o f eroded soil. These physical stressors are
often accompanied by chemical stressors associated
with the land conversion itself, chemical stressors
from the new use o f the converted land, or sedimentborne chemical inputs. In the case of conversion o f
BHRE to pasture or the use of BHRE for grazing,
forage, and shelter of livestock, the physical stressors
are also associated with biological stressors.
Conversion of B H RE to other vegetation is possible
without drainage modifications if the replacement
vegetation is able to tolerate the poorly drained conditions and periodic flooding of BHRE. T h e two primary types of vegetation conversion are to convert to
pine plantations or to convert to wet pastures. Native
pine species, such as slash pine (Pinus eUiottii Englem.)
and spruce pine (P. glabra Walt.), are well adapted to
riparian conditions. Slash pine habitat is described as
"low ground, hammocks, swamps, and along streams"
(Harrar and H a r r a r 1962). Spruce pine occurs in similar habitats and has been observed to occur on banks
o f the major rivers that originate in the TVU. Conversion to pastures without drainage modifications is often accompanied by an increase in Juncus spp. and
other wetland-tolerant plants.
T h e removal and drainage of BHRE is perhaps the
ultimate stressor, with complete loss of the WQF. T h e
physical stressor is typically accompanied by direct
247
input of chemical stressors in agricultural practices.
For purposes o f this conceptual model, these direct
chemical inputs will not be considered chemical stressors for the physically converted systems but will be
considered as part of the exposure scenario for downstream systems. Direct conversion o f BHRE to cropland is generally possible only with accompanying
drainage and soil modifications. Typical practices in
the T V U are to expand fields for use o f larger center
pivot irrigation systems. First- and second-order
stream channels are converted to ditches fed by lateral
lines o f subsurface perforated pipe (tile drainage).
Almost all artificial drainage in the T V U is for this
type of field expansion rather than large-scale drainage of bottomlands along major rivers (Soil Conservation Service, Tifton County personal communication). T h e USDA has occasionally summarized data
on drainage of agricultural lands in special Census o f
Agriculture reports. T h e most recent data available
(1978) show that about 173,000 ha o f farmland was
drained (Table 2). This is about 14% o f the total 1987
cropland.
Long-term sediment deposition has apparently altered the nature of riparian soils in the TVU. Anecdotal accounts suggest that the presettlement streams
had better defined, less braided channel systems than
presently exist on most lower-order streams (Harper
1906, Lowrance and others 1948a). Estimates o f sediment deposition rates indicated an average o f 35-52
mg/ha/yr for 1880 through 1979 (Lowrance and others 1984a, 1986). This would indicate an average of
22-33 cm of deposition in this time period. It is not
possible to determine whether this deposition has increased or decreased the area o f riparian wetlands,
although the net effect of the sediment delivery to the
channel systems should be to increase hydrologic retention.
Biological stressors. Although physical and chemical
stressors are emphasized in the Risk Assessment
Framework (RAF 1992) and are the focus o f this conceptual model, a special case o f biological stressor-animal production units such as cattle and swine--will
also be considered. This is necessary because o f the
use o f BHRE for grazing, forage, and shelter o f livestock (especially cattle, but some swine). Livestock can
be a major factor in transfer of nutrients in agricultural landscapes (Woodmansee and Adamsen 1983).
In the TVU, livestock represen~ a mobile means o f
introducing chemical, biological, and physical stressors into the riparian environment. Primary potential
impacts o f the introduction o f livestock in BHRE are
the loss and/or damage of native vegetation, direct
introduction of waste products into the stream chan-
248
R. Lowrance and G. Vellidis
nel system, and compaction of wetland soils through
trampling.
Ecosystems Potentially at Risk
T h e wetland ecosystems potentially at risk have
been described by Wharton (1978) as blackwater
(nonalluvial) river systems. According to Wharton's
description, this includes both the blackwater river
and swamp system and the tributary streams described as blackwater branch or creek swamps. This
description will also include a third ecosystem type
described by Wharton, the bay swamp, which in the
T V U is hydrologically connected with the blackwater
river systems.
Wharton (1978) describes these river and swamp
systems as having narrow floodplains and lacking extensive tracts of bottomland hardwoods. T h e Withlacoochee, Little, and Alapaha rivers inundate their
fairly narrow floodplains for long periods of time
(Faircloth 1971). Swamp black gum [Nyssa sylvatica
var. biflora (Walt.) Sarg.] dominates the floodplains
where water movement is restricted. Water tupelo
(Nyssa aquatica L. ) and bald cypress [Taxodium distichum (L.) Rich.] tend to grow in more open water with
more circulation. T h e riparian areas along these
streams are sometimes dominated by low areas of
slash pine. An extreme example of this is along the
Satilla River, where an informal survey in 1971
showed that about 320 of the total 365 km of channel
were bordered by pine lowlands (Wharton 1978).
T h e tributaries o f the blackwater rivers originating
in the T V U are described as blackwater branch or
creek swamps (Wharton 1978). These broadleaf tree
and shrub communities occur as bands of vegetation
on moist organic soils along small streams. H a r p e r
(1906) described these branch swamps in the T V U
and indicated that about 50% of the woody plants in the
branch swamps were evergreen or tardily deciduous.
A third ecosystem type associated with the BHRE
has been described by Wharton (1978) as the bay
swamp. These ecosystems are dominated by sweet bay
(Magnolia virginiana L.) and loblolly bay [Gordonia lasianthus (L.) Ellis]. These ecosystems occur at the
edges of floodplains where there is groundwater
seepage occurring above aquicludes such as the plinthic material or Hawthorne formation. T h e bay
swamps have up to 2 m of accumulated peat with an
organic matter content of 41%-98% (Wharton 1978).
Exposure Scenarios
Establishing exposure scenarios may be the most
difficult and controversial phase of any ecological risk
assessment. Although it may be fairly easy to identify
potential stressors, it is exceedingly difficult to foresee
corresponding pragmatic exposure scenarios because
of their complexity. A myriad of political, economic,
and social variables ultimately determine the extent to
which stressors operate on the ecosystems at risk. For
example, let us examine logging of bottomland hardwood forests as a stressor. Data on wetland forest
logged annually can be compiled. T h e questions then
are how this historical data can be used to project
future trends and what other factors should be considered in the projections. Future d e m a n d for hardwood fiber, timber tax laws, the Wetland Reserve Program, and definition of jurisdictional wetlands are
only some of the factors that play a pivotal role in
determining the extent of future logging.
Because the task o f predicting real-life exposure
scenarios requires intricate politicosocioeconomic
models that are beyond the scope o f an ecological risk
assessment, it may be more relevant to evaluate simplified exposure scenarios. These scenarios would involve quantifying exposure as a function of magnitude and/or areal extent o f the stressor. In the case of
bottomland hardwood logging, potential exposure
scenarios could be the clear-cutting of 10%, 25%, and
50% of all remaining bottomland hardwood forests in
the TVU, or more sophisticated scenarios, such as the
impact of the redelineation of jurisdictional wetlands
on logging of bottomland hardwood forests, could be
evaluated.
In this manner, ecological risk could be assessed
lor a series of exposure scenarios that would theoretically bracket actual events. Exposure scenarios resulting from specific policy or management decisions can
be best predicted by risk managers with local knowledge o f the ecosystems at risk. T h e ecological effects
analyses developed by risk assessors can be most appropriately utilized by risk managers to determine the
impact o f exposure scenarios resulting from specific
policy decisions.
Exposure scenarios for the movement of chemical,
physical, and biological stressors from agricultural
lands into BHRE require methods to determine the
juxtaposition of agricultural uses with BHRE and the
ability to predict or model the effects of nonpoint
source pollutants on the WQF. It is possible that existing spatially distributed data bases will be adequate
for delineating these relationships in representative
watersheds of the TVU. Digital National Wetlands
Inventory maps can be overlain on digital line graphs
of 7.5-min quad sheets to determine the NWI wetlands associated with known stream channels. These
data can then be combined with the Georgia Landcover Data Base developed for the Georgia Depart-
Model to Assess Ecological Risk
Table 4.
249
Agricultural chemical inputs to Tifton-Vidalia Upland Farms a
Agricultural chemical
Commercial fertilizer
On cropland
On pastureland &
rangeland
Lime
Insecticides
Nematicides
Fungicides, etc.
Herbicides
Defoliants, growth
regulators, etc.
(farms)
(ha)
(farms)
(ha)
(farms)
(ha)
(farms)
(ha)
(Mg)
(farms)
(ha)
(farms)
(ha)
(farms)
(ha)
(farms)
(ha)
(farms)
(ha)
Georgia
TVU
Percent
of
Georgiab
28,762
1,298,753
21,993
1,076,702
11,040
222,051
9,002
229,203
494,140
11,459
552,647
3,494
112,605
5,217
176,755
14,402
789,162
1,837
85,376
12,333
706,281
11,291
645,191
3,098
61,091
3,783
109,771
228,369
7,534
359,697
2,541
71,321
3,096
97,867
7,379
463,091
1,293
63,219
42.9
54.4
51.3
59.9
28.1
27.5
42.0
47.9
46.2
65.7
65.1
72.7
63.3
59.3
55.4
51.2
58.7
70.4
74.0
Percent
of TVU
land area
Percent
of TVU
farmland
14.0
34.4
12.7
31.4
1.2
3.0
2.2
5.3
7.1
17.5
1.4
3.5
1.9
4.8
9.1
22.5
1.2
3.1
Percent
of TVU
farms
78.4
71.8
19.7
24.1
47.9
16.2
19.7
46.9
8.2
"Data are for 1987and were complied from the 1987Census of Agriculture (1989).
hProportionof Georgia included in the Tifton-VidaliaUpland for each chemical.
ment o f Natural Resources. This data base delineates
15 landcover classes, including pasture and cultivated
land. Data on acreages receiving agricultural chemicals are available (Table 4) as are annual records of
crop rotations on a field-by-field basis that are reported by farmers to the USDA-ASCS and mapped
on aerial photographs for each county. Records o f
actual chemical inputs are available for limited releases beginning in 1993, but state recommendations
are representative of average application rates. Once
the juxtaposition of agricultural land and BHRE has
been determined, a large n u m b e r of exposure scenarios can be postulated based on the potential stressors
defined above.
Chemical stressor exposure scenarios. Chemical stressors in the T V U can be generally grouped into two
categories: those resulting from agricultural and forestry practices and those resulting from urban/
industrial sources. Because the TVU is primarily an
agricultural region, much of which is adjacent to riparian forest wetlands along first-, second-, and thirdorder streams, the stressors of greatest importance
are plant nutrients and pesticides leaving agricultural
production systems and entering BHRE. T h e movement o f nonpoint source pollutants into BHRE depends on hydrologic connections, which can be determined using the topographic and hydrographic data.
T h e extent to which the WQF of BHRE located
adjacent to agricultural production sites is affected by
these pollutants is not well documented. T h e relative
impact o f various exposure scenarios can best be evaluated by linking a series o f models. Results of GIS
manipulations reflecting chemical inputs to agricultural upland fields in selected representative watersheds can be used to initialize transport models such
as CREAMS (Knisel 1980) and GLEAMS (Leonard
and others 1987). T h e transport models can be used
to provide relative concentrations o f pollutants leaving the fields in surface r u n o f f or shallow groundwater flow. These edge-of-field loadings provide input
to wetland models such as the Riparian Ecosystem
Management Model (REMM) (Altier andothers 1994)
to simulate fate and transport through BHRE.
Although landscape applications o f some o f these
detailed, process-based models have been attempted
(Stallings and others 1992), linking them with spatially distributed data bases through a GIS will be difficult for entire watersheds. Models based on applications and modifications of existing GIS capabilities
are more likely to yield useful results for watershedscale simulations. Models of these sorts, using simple
functional relationships for nonpoint source pollutants and sinks for pollutants have been developed by
Kesner and Meetemeyer (1989) and Levine and oth-
250
R. Lowrance and G. Vellidis
ers (1993). These models link either statistical or
mass-balance modeling of the nutrient and/or sediment delivery with the spatial arrangement of soil, vegetation, and slope properties that control transport.
It must be noted that simulation results, at best,
provide relative comparisons between exposure scenarios and should not be used to predict concentrations o f pollutants at any stage. T h e modeling process
can be calibrated with field data accumulated during
the quantitative phase o f a risk analysis study.
Drift and improper application. Quantifying pollutant
loads from spray drift and improper application o f
agrichemicals is an exercise in probability theory and
beyond the scope o f this conceptual model. Nevertheless, the ecological effect of these activities on BHRE
may be very real and should be investigated with some
simplified exposure scenarios. T h r o u g h simulation,
the relative impact of a range of pesticide loads applied to BHRE vis spray drift can be examined. Similarly, scenarios o f improper agrichemical applications
on agricultural production sites can be simulated and
the resulting edge o f field loadings used to initialize
the BHRE models.
Animal waste nutrients. Animal confinement facilities
produce tremendous quantities of animal m a n u r e
with high concentrations of nutrients and, in some
instances, heavy metals. A herd of 1000 dairy cattle
will produce waste equivalent to a city o f 10,000 (Hubbard and Lowrance 1994). Based on USDA-SCS recommendations over the past two decades, farmers
constructed sorage lagoons to receive liquid m a n u r e
resulting from flush cleaning o f confinement facilities
and r u n o f f originating on feedlots. Many o f these
lagoons have now reached their design capacities and
waste management has reached a crisis stage. Some
lagoons regularly overflow during intense or prolonged rainfall events discharging untreated or marginally treated manure with concentrations ranging as
high as 200 mg N/liter-~ and 50 mg P/liter-i directly
into adjacent BHRE and streams. An increasingly
popular solution is to land apply the liquid manure
(typically containing less than 3% solids) through an
irrigation system. Although research is currently underway to establish nonpolluting application rates,
many land application systems are already in operation. Application rates as high as 1000 kg N/haJyr are
not unusual. T h e percentage of applied N lost to ammonia volatilization and denitrification has not yet
been well quantified for these systems. Nevertheless,
u n d e r most cropping systems, this rate provides close
to double the annual N required by the crops.
Exposure scenarios for land-applied animal waste
can be evaluated as described above for conventional
agricultural chemical stressors. Nutrient discharges
from lagoons can generally be considered point
sources. Chemical stressors introduced into the envir o n m e n t as point sources (effluent from sewage treatment plants, pesticide formulation plants, and fertilizer distribution centers) generally enter the
streamflow o f larger-order streams without passing
through a BHRE. Thus, the impact of these pollutants on BHRE and the effect o f BHRE on their assimilation and/or degradation is limited in the T V U landscape. Loadings from point sources, however, can
skew or mask the effect of stressors on the WQF of
B HRE du~ing watershed-scale evaluations. Care must
be taken to account for these sources during any
quantitative risk analysis.
Physical stressor exposure scenarios. For the purpose
o f this conceptual model, physical stressors deal primarily with the conversion o f BHRE to other land
uses, whether this is clearing the natural vegetation
for pasture, conventional agriculture, urban development, or drainage. In the TVU, BHRE losses have
been traditionally related to agriculture. Recent data
show that the percentage of farmland in the landscape is decreasing and that conversion due to agricultural uses may decrease.
Conversion to or from farmland. In the 1987 Census of
Agriculture (1989), 5,075,702 acres were designated
as farmland within the TVU, down 21% from 1978.
During the same period, harvested cropland decreased by 30.5%. This decrease reflects the well-documented financial difficulties faced by farmers
throughout the United States during the past decade.
Agricultural economists are not sure that this trend
will continue into the 1990s but generally agree that,
barring dramatic changes in farm policy, farmland
will not increase in the near future. Changes in the
definition of jurisdictional wetlands, application of
the Wetland Reserve Program, and changes in pricesupport programs, particularly peanuts---the most
important agricultural crop in the TVU, are key issues that are likely to impact farming patterns and
BHRE.
During the same period, the TVU saw a consolidation of farmland and some conversion o f BHRE to
agricultural land---dependent in part on the need for
expansion of adjacent fields. Because the T V U landscape is dominated by upland agricultural areas interlaced with riparian forests along low-order stream
channels, fields are commonly enlarged by clearing
the forest and ditching the stream channel. This can
be done for a variety of reasons---to accommodate
irrigation systems or simply to make machinery operation easier. Fields of a certain size and in a certain
topographic position are more likely to expand for
the use of large center pivots. Data on the n u m b e r and
Model to Assess Ecological Risk
size of fields irrigated by large-scale fixed irrigation
systems can be extrapolated from information available through the University o f Georgia Cooperative
Extension Service. Agricultural census data indicate a
35% increase in irrigated acreage in the T V U between
1978 and 1987.
BHRE that are likely candidates for conversion to
accommodate larger fields can be identified from the
GIS data base, especially if property boundaries and
ownership are not considered (Vellidis and others
1987). Risks associated with these conversion scenarios can be evaluated on both a site-scale and a watershed-scale level.
Logging. Several hardwood logging and processing
operations are located within the TVU. T h e extent o f
their logging operations is not well documented,
partly because o f the reluctance o f the loggers to
openly discuss their operations. Data on annual harvest of bottomland hardwoods are not available. Anecdotal evidence indicates that marketable stands o f
bottomland hardwoods are being depleted in the
TVU.
Management practices during logging operations
vary greatly from one operator to another. It is common to log straight across a stream, completely removing all marketable trees and in the process severely disturbing the stream channel. Recommended
Best Management Practices (BMPs) are to minimize
disturbance o f the channel itself by establishing a limited n u m b e r o f crossing points. Logged areas are generally either replanted after harvest with pines or are
allowed to regenerate naturally. Logical exposure scenarios would be to compare the ecological effect of
these two management practices. Additional scenarios
that can be postulated include evaluating the effect o f
leaving a band of trees o n either side o f a stream,
reforestation following clear-cutting, and thinning o f
BHRE. Data from ongoing research by the University
o f Georgia Biological & Agricultural Engineering Department and the USDA-ARS Southeast Watershed
Research Laboratory at two experimental sites near
Tifton can be extrapolated to simulate the ecological
effect o f the selected exposure scenarios on similar
ecosystems in the TVU.
Removal and drainage. Artificial drainage in the T V U
is almost entirely associated with field expansion (discussed earlier) rather than large-scale drainage o f bottomlands along rivers. Changes in farm ecinomics,
redelineation o f jurisdictional wetlands, or other factors may influence this trend. T h e removal and drainage o f BHRE results in the complete loss of the WQF.
In this case, edge-of-field (or "edge-of-parking-lot" in
the case o f removal and drainage for development)
loadings are a direct source of contamination for re-
251
ceiving waters. Exposure scenarios for removal and
drainage are probably most appropriately evaluated
on a watershed scale.
Conversion to wet pasture. Conversion o f B H R E to
other vegetation, particularly pastures or pine plantations, without drainage modifications is a fairly common practice along lower-order streams. Conclusive
data on the effect o f such conversions on the WQF o f
riparian areas have not yet been published. Research
currently in progress in England indicates that wet
pastures (without foraging livestock) are effective
sinks of nutrients (Haycock personal communication).
Data on shallow groundwater quality in wet pastures
o f the T V U are available (Lowrance and others 1983).
These ecosystems can best be considered by postulating that their current WQF is x% o f their original
WQF on the premise that original designates undisturbed BHRE. For example, a B H R E converted to a
mature pine plantation might retain close to 100% o f
its original undisturbed WQF, whereas a wet pasture
might retain only 50%.
Biologicalstressor exposure scenarios. Livestock can be
a major factor in transfer o f nutrients, trampling and
removal o f vegetation, compaction of soils, and trampling of streambanks in agricultural landscapes and
adjoining BHRE. It is quite common to see livestock
standing or wallowing in stream channels during the
hottest parts o f the day. These activities result in increased nutrient loads, reduced infiltration coupled
with higher erosion rates, and possibly lower overall
nutrient uptake by the remaining vegetation. Exposure scenarios can be evaluated in terms o f an increase
or decrease in acreage devoted to this activity and in
terms o f an increase or decrease in livestock concentration in grazing areas. No directly applicable data on
the effect o f livestock presence on the WQF or BHRE
are available.
Measurement End Points
T h e assessment end point is the water quality function o f bottomland hardwood riparian ecosystems.
Ultimately, we would like to know the risk to this
function associated with the potential stressors operating through a range o f exposure scenarios. One
paradox o f the risk assessment is that the nonpoint
source pollution retention aspect o f the WQF probably becomes more important to downstream water
quality as land use in a watershed intensifies, that is, as
the function becomes more at risk. It is not yet known
whether the WQF or BHRE is linearly related to the
inputs and whether there is a threshold beyond which
further stressors lead to a nonlinear decline in the
WQF. Within limits, BHRE should be able to respond
to increases in nutrient nonpoint source pollutants by
252
R. Lowrance and G. Vellidis
increasing nutrient cycling rates. Evidence for this is
found in higher denitrification rates for riparian systems receiving higher nitrogen, water, and organic
matter subsidies (Lowrance unpublished). Increased
growth by vegetation in response to nutrient subsidies
is a n o t h e r example of the WQF potentially adapting
to the stressor (Fail and others 1986). O t h e r stressors,
such as pesticides and sediment, are less likely to cause
adaptive changes in the BHRE, although it is possible
that certain insecticides could actually benefit vegetation and that sediment-borne nutrients could stimulate either microbial or vegetation processes. Stressors
such as conversion and livestock access are less ambiguous and can be seen as generally producing a deterioration in the WQF. T h e major difficulty in assessing
the effects of these stressors is in gauging the watershed- or landscape-level effects of converting or providing animal access to individual stands o f BHRE.
It is necessary to define m e a s u r e m e n t end points
for the WQF at two levels---individual sites and watersheds. Measurement end points are necessary at these
two levels for at least three reasons: (1) stressors can
operate on B H R E at either the individual stand level
or at the wateshed level and the functions can be impaired at either of these scales; (2) certain land uses,
such as forestry and agriculture, will come u n d e r increasing scrutiny in the future to determine water
quality effects of m a n a g e m e n t o f individual ownership units such as farms or forest tracts; and (3) the
risk to functions due to stressors at individual sites
might be minor, but the integrated risk at the watershed scale might be significant. We will discuss the
development of m e a s u r e m e n t end points for these
two scales (Table 5).
At both the site and watershed level, each measurement end point may be useful for assessing the risk
f r o m more than one stressor. Conversely, stressors
will generally be assessed using m o r e than a single
m e a s u r e m e n t endpoint. This integration of measurement end points is explicit in Figure 5, as in the integration o f stressors. It is implied in Figure 5 that measurement end points may provide information about
the risks associated with m o r e than one stressor.
Measurement end points should generally be objective parameters that relate to the WQF at one of the
scales. T h e problem with application o f these measurement end points to the risk analysis is that objective measures require evaluation criteria to determine
if values of the m e a s u r e m e n t end points imply a risk
to the functions. T h e evaluation criteria can only be
provided by the use of reference sites/watersheds
where it can be shown that the W Q F o f B H R E is not
impaired. T h e selection o f reference sites/watersheds
Table 5. Measurement endpoints for the
Tifton-Vidalia BHRE ecological risk assessment
Individual site scale
Hydrology/water quality
Infiltration rate for rainfall
Runoff rates
Depth to water table
Duration and depth of inundation
Agrichemical concentrations in runoff
Agrichemical concentrations in shallow groundwater
Nitrate/Chloride ratios
Soil
Denitr~cation potential
Redox potential
Pesticide residues
Vegetation
Nitrate reductase level
Woody biomass accumulation rate
Diversity and phenology
Pesticide residues
Watershed scale
Hydrologic characteristics
Percent and timing of storm events relative to
base flow
Average slope length from edge of field to
stream channel
Hydrology/pollutant relationships
Changes in nutrient concentrations in storm
events
Location of point sources relative to BHRE
Overall agrichemical loading of a watershed
Overall sediment loading of a watershed
Land-use relationships
Area and age of converted BHRE
Continuity of BHRE related to the stream
channel network
Length of interface between cropland or pasture
land and BHRE
will be complicated by adaptation to stressors discussed above. For the ecological risk assessment, reference sites/watersheds should include both pristine
and healthy BHRE. T h e pristine sites/watersheds
would be those with the least h u m a n impact, such as
B H R E in completely forested watersheds. T h e
healthy sites/watersheds would be those representative o f the land use in the T V U but would be identified as having B H R E with intact WQF.
Site-scale measurement end points. Site-scale measurement end points for the W Q F fall into three general
categories: soil, vegetation, and hydrology. T h e y have
been selected based on the potential for relatively
short-term measurements that will be indicative of the
WQF o f a site. Candidates for these m e a s u r e m e n t end
points would include: denitrification potential or rate,
redox potential, soil pesticide residues, nitrate reductase levels in vegetation, woody biomass accumulation
M o d e l to A s s e s s Ecological Risk
~ :~i :'~!i:ii~li:iii~~/i~;
~~:~i~!
~~ii~i
' ii::~
Nr
r
::!!~ i
~'
I
EL:
]
Fig. 5. Linearized methodology for assessing ecological risk
to bottomland hardwood riparian ecosystems in the TiftonVidalia Upland.
253
rate, vegetation diversity and phenology, vegetation
pesticide residues, infiltration rate for rainfall and
runoff, depth to water table, and duration and depth
of inundation (Table 5).
As with any set of measurement end points, there
will be a wide range in the a m o u n t o f data available
and the current state of knowledge o f how a measurement end point relates to the assessment end point.
For instance, denitrification rates in BHRE of the
T V U have been measured u n d e r a variety o f conditions and with a variety of stressors present (Ambus
and Lowrance 1991, Hendrickson 1981, Lowrance
1992, and other ongoing research). Redox potential
has been shown to be one of the major controls on
whether denitification can occur in soils of BHRE
(Gambrell and others 1975). Vegetation uptake is also
an important means of sequestering plant nutrients in
BHRE, which might otherwise lead to downstream
eutrophication (Fail and others 1986). Runoff, rainfall, and associated pollutants that infiltrate in BHRE
is subject to pollution reduction processes due to hydrologic retention in the system (Shirmohammadi
and others 1986). Nitrate/chloride ratios characteristically decrease rapidly in BHRE with substantial nitrate removal (Lowrance 1992). Some measurement
end points have potential but are more poorly understand. Data on pesticide residues in BHRE soils and
vegetation are probably nonexistent for the TVU. It is
not well understood how vegetation diversity and
phenology might affect the WQF. Although many
woody plants respond to nitrate subsidies by increased levels o f nitrate reductase (Smirnoff and others 1984), nothing is known about this response in
BHRE vegetation.
Watershed-scale measurement end points. These end
points fall into three main categories: hydrologic
characteristics, hydrology/pollutant relationships, and
land-use relationships. T h e y have been selected based
on the potential for using existing data bases on land
use, land cover, soils, etc., and based on existing
knowledge of hydrologic and hydrology/pollutant relationships in the TVU. Hydrologic, nutrient, and
sediment transport data bases for Little River Research Watershed (LRRW) (Batten 1980, Lowrance
and Leonard 1988)collected by USDA-Agricultural
Research Service sample a range o f watershed land
uses from about 30% agricultural to about 70% agricultural. These data bases are available for varying
time periods from 1966 through the present for watersheds ranging in size from 260 ha to 33,400 ha.
Some detailed land-use data are also available for
some watersheds. Data from these watersheds would
be used to represent healthy agricultural watersheds
254
R. Lowrance and G. Vellidis
with a range o f agricultural land uses. Other existing
data bases on water quality and hydrology in the T V U
consist primarily of limited streamflow and stream
water quality data collected by the US Geological Survey and the Georgia Department of Natural Resources.
Candidates for the watershed-scale measurement
end points are: percent and timing o f storm events
relative to baseflow; changes in nutrient concentrations in storm events, area and age o f converted
BHRE, average slope length from edge of field to
stream channel, continuity of BHRE related to the
stream channel network; location of point sources relative to BHRE, length of interface between cropland
or pasture land and BHRE, and overall agricultural
loading to a watershed. Data exist for most of these
watershed-scale parameters from studies conducted
on the LRRW.
Although the WQF of BHRE is only one determinant o f regional water quality, studies in the T V U
show it to be a critical factor. T h e r e f o r e measurement
end points that relate to regional water quality should
be useful in assessing the WQF on a regional or basin
scale. Hydrologic relationships and hydrology/pollutant relationships are critical to assessing the regional
water quality impact of loss o f the WQF of BHRE.
Certain relationships can be attributed to the loss o f
the WQF, particularly increases inflow and increases
in dissolved nutrients such as nitrogen and phosphorus species during steamflow events, attributable to
short-circuiting of the nutrient retention in BHRE
(Lowrance and Leonard 1988). Using regional water
quality measurements to infer the WQF of BHRE
requires both land-use and landscape data for a number o f watersheds but will also require information on
point sources of pollutants. T h e ideal situation to address the regional WQF of BHRE would be watersheds with similar land uses, no point sources o f pollution, and differing areal extent o f BHRE or BHRE
with distinctly different stressors. This situation probably does not exist in the TVU and certainly does not
exist for gauged watersheds with water quality monitoring.
Data for landscape relationships for watersheds in
the T V U can be developed from existing aerial photographs and data bases such as those discussed u n d e r
exposure scenarios, along with land use surveys of
watersheds other than the LRRW. Data on hydrologic
relationships, and hydrology/nutrient transport relationships are only available from the LRRW, although
data may become available from two other mixed
land-use watersheds as part of the US Geological Survey National Water Quality Assessment (NAWQA)
Program. Data on pesticide transport on LRRW
would need to be obtained as part of a quantitative
risk assessment, although limited data may be obtained through NAWQA.
Risk Analysis and Characterization
A risk assessment study may be probabilistic, deterministic, or qualitative in the approach taken to express ecological risk. Regardless o f the approach, a
basic structure and starting principles are necessary.
Figure 5 presents a linearized methodology for assessing ecological risk to BHRE in the TVU. T h e flowchart is loosely based on EPA's Frameworkfor Ecological
Risk Assessment (RAF 1992) but is organized in a stepwise fashion to show a logical sequence of events during a study. T h e iterative structure signifies that a
given action is incremented i times. For example, i key
stressors may be identified in the problem formulation phase. Each identified stressor warrants j exposure scenarios, each o f which may have a series o f
ecological effects that need to be evaluated individually. Each ecological effect is evaluated with a series of
measurement endpoints. This does not preclude the
concurrent use of a given measurement end point for
assessing several ecological effects. Furthermore,
measurement end points may range from field to watershed scale.
T h e measurement end points used for each ecological effect are integrated to provide a comprehensive
assessment of the ecological effect of the given stressor on the identified ecosystem at risk. T h e impact of
the stressor on the ecosystem functions can then be
prioritized before the next ecological effect is considered.
T h e integration step shown following the completion of each loop is intended to provide a synthesis of
assessments made during each iteration o f the loop.
Following the integration o f all identified ecological
effects for a given stressor, the impact on the ecosystem functions is revisited to permit a more global prioritization of risk than was done previously. T h e logical conclusion o f the risk assessment study provides a
prioritized list o f stressors. T h e prioritized list o f stressots is then used to develop management strategies
for reducing risk.
It is, of course, recognized that during a quantitative risk assessment study, many of the items presented here in iterative form in the flowchart will be
taking place concurrently rather than sequentially. It
is also recognized that there is likely to be significant
overlap between ecological effects and measurement
end points, which adds another level of complexity
Model to Assess Ecological Risk
Table 6.
255
Specific example of an ecological risk assessment in the TVU
Stressor
Exposure scenario
Ecosystem at risk
Potential ecological effects ( 9 ) and measurement
endpoints (o)
that has not been addressed in Figure 5. Nevertheless,
the linearized approach does provide a useful stepwise guide.
As an illustration, consider the ecological risk presented by a specific stressor, in this case, logging/clearcutting o f BHRE. A possible exposure scenario could
be the clear-cutting o f 25% o f all remaining bottomland hardwood forests in the T V U over a specified
period o f time. T h e ecosystem at risk would then be
defined generally as bottomland hardwood forests
along blackwater river and swamp systems and specifically based on either some knowledge o f where logging would occur or assumptions about distribution.
Table 6 presents three potential ecological effects
likely to result f r o m the stressor operating through
the given exposure scenario on the ecosystems at risk.
It also presents a series o f m e a s u r e m e n t end points
that can be used to quantify the ecological effects.
Because only one stressor and only one exposure scenario are being assessed, the stressor and exposure
scenario steps do not need to be incremented. In this
example the iterative structure is not active until the
ecological effects step, at which point each of the three
predefined ecological effects are evaluated sequentially.
Logglng/clear-cutting
Clear-cutting of 25% of all remaining bottomland
hardwood forests in the TVU
Bottomland hardwood forests along blackwater river and
swamp systems in the TVU
9 Reduction in overall attenuation and removal of
agrichemicals from shallow ground water and surface
runoff flowing through logged riparian zones
9 agrichemicai concentrations in runoff
9 agrichemical concentrations in shallow groundwater
9 soil denitrification potential
9 soil agrichemical concentrations and residues
9 changes in nutrient concentrations in storm events at
watershed outlets
9 Increased runoff rates, reduced time to concentration,
increased discharge rates, and subsequent high peak
flow and increased potential for stream and channel
erosion
9 infiltration rate for rainfall
9 runoff rates
9 duration and depth of inundation
9 percent and timing of storm hydrologic
characteristics relative to base flow
9 Increased sedimentation in streams draining logged
areas as the erosion rate in logged riparian zones
increases and runoff and sediment trapping capacity of
the riparian zones decreases
9 overall sediment loading of the watershed streams
and channels
T h e impact of each ecological effect on the W Q F o f
the ecosystem at risk is d e t e r m i n e d by using the appropriate m e a s u r e m e n t end points such as concentrations o f agrichemicals in runoff, shallow groundwater, and streamflow at watershed outlets; soil
denitrification potential; and soil agrichemical
concentrations. These must be integrated in the context o f the first potential ecological effect to determine if logging/clear-cutting will result in reduced agrichemical attenuation.
Integration is a broad step that may include any
combination o f methods to synthesize the data and is
likely to include both quantitative and qualitative
j u d g m e n t s based u p o n the evaluators' experiences
and understanding o f the ecosystem at risk. As a result, integration has frequently been an area o f contention in past ecological risk assessments. T h e mechanism used to integrate the m e a s u r e m e n t end points
for each o f the three potential ecological effects is
likely to differ based on the reliability o f field measurements, length o f the data set, and understanding
o f the governing principles o f the ecosystem function
being affected. T h e same analysis is repeated for each
of the other two ecological effects (Table 6). T h e three
ecological effects are then integrated to d e t e r m i n e
256
R. Lowrance and G. Vellidis
their cumulative effect, and ultimately the effect o f
the stressor being evaluated, on the WQF o f the ecosystems at risk. In risk assessments that include several
stressors, a prioritization of cumulative effects associated with specific stressors should also be clone at this
stage. In the event that several stressors have been
identified, the above procedure is repeated for each
stressor. During integration, uncertainty analysis
must be performed to establish some level of confidence in the conclusions. I f it is determined that the
WQF of the ecosystem at risk is affected by the given
ecological effect, some measure o f importance or priority with respect to the other associated ecological
effects and stressors should be attached to the ecological effect.
Cumulative impacts o f anthropogenic activities on
bottomland hardwood wetland ecosystems have been
explored in detail in a series of USEPA-sponsored
workshops (Gosselink and othes 1990a). Cumulative
impacts are defined by the Federal Council on Environmental Quality as "the impact on the environment
which results from the incremental impact of the action when added to other past, present, and reasonably foreseeable future actions. Cumulative impacts
can result from individually minor but collectively significant actions taking place over a period o f time."
(40CFR part 1508.7 and 1508.8, cited in Harris and
Gosselink 1990). Although bolstered by developments
in ecological theory and "stress ecology," the actual
assessments o f cumulative impact have been j u d g e d
marginally effective by the authors providing the
most thorough evaluation of the field (Harris and
Gosselink 1990).
Although cumulative impacts are known to be important in affecting the functions of wetlands, including the WQF, only limited success has been reported
in assessing cumulative impact. Gosselink and others
(1990b) reported on a best-professional-judgement
synthesis of effects of a number of generic management activities (channel works, clearing, draining, impoundment, isolation/constriction) on the hydrologic
functions of bottomland hardwood ecosystems. Each
activity incorporates numerous stressors and the impact on the hydrologic function is assessed as high,
medium, low, or none. Flood mediation, as a hydrologic function, is assessed in more detail. These hydrologic assessment models summarize the impacts in
a qualitative way by integrating the various activities in
flowcharts and tables to provide a first approximation
of cumulative impacts. A series of conceptual assessment models for nonpoint source pollution control
functions o f bottomland hardwood ecosystems was
also developed as part of the same workshops (Scott
and others 1990). These assessment models for nonpoint pollution control did not specifically address
cumulative impacts. A similar conceptual assessment
model for nitrate removal by riparian ecosystems was
developed by Pionke and Lowrance (1991).
Quantitative assessment of cumulative risk may go
beyond both the limits o f our empirical knowledge of
BHRE and the limits o f our ability to predict risk
based on functional principles. It is likely that the
prioritization o f risks will provide a more useful
framew_ork. In some cases, the need to assess cumulative risf~ will be superseded by obvious loss o f the WQF
due to a limited set of activities. T h e assessment of
cumultative risk should be o f considerable importance in a landscape such as the T V U where BHRE do
or can perform a crucial water quality function.
T h e final products of the risk assessment should
include an ecological risk summary, total uncertainty
analysis, and if applicable, a prioritized list of stressors
(Figure 5). These products can then be used by risk
managers to develop management strategies for reducing risk.
Acknowledgments
This research is supported by Funds from the
United States Environmental Protection Agency-Environmental Research Laboratory (Corvallis), Hatch
and State funds allocated to the Georgia Agricultural
Experiment Stations, and funds allocated to the
USDA-ARS Southeast Watershed Research Laboratory.
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