REVIEWS & ANALYSES
Water Resources and Land Use and Cover in a Humid Region:
The Southeastern United States
R. Chelsea Nagy,* B. Graeme Lockaby, Brian Helms, Latif Kalin, and Denise Stoeckel
It is widely recognized that forest and water resources
are intricately linked. Globally, changes in forest cover to
accommodate agriculture and urban development introduce
additional challenges for water management. The U.S.
Southeast typifies this global trend as predictions of land-use
change and population growth suggest increased pressure on
water resources in coming years. Close attention has long
been paid to interactions between people and water in arid
regions; however, based on information from regions such as
the Southeast, it is evident that much greater focus is required
to sustain a high-quality water supply in humid areas as well. To
that end, we review hydrological, physicochemical, biological,
and human and environmental health responses to conversion
of forests to agriculture and urban land uses in the Southeast.
Commonly, forest removal leads to increased stream sediment
and nutrients, more variable flow, altered habitat and stream
and riparian communities, and increased risk of human health
effects. Although indicators such as the percentage of impervious
cover signify overall watershed alteration, the threshold to
disturbance, or the point at which effects can been observed
in stream and riparian parameters, can be quite low and often
varies with physiographic conditions. In addition to current
land use, historical practices can greatly influence current water
quality. General inferences of this study may extend to many
humid regions concerning climate, environmental thresholds,
and the causes and nature of effects.
Copyright © 2011 by the American Society of Agronomy, Crop Science
Society of America, and Soil Science Society of America. All rights
reserved. No part of this periodical may be reproduced or transmitted
in any form or by any means, electronic or mechanical, including photocopying, recording, or any information storage and retrieval system,
without permission in writing from the publisher.
J. Environ. Qual. 40:867–878 (2011)
doi:10.2134/jeq2010.0365
Posted online 23 Feb. 2011.
Received 18 Aug. 2010.
*Corresponding author ([email protected]).
© ASA, CSSA, SSSA
5585 Guilford Rd., Madison, WI 53711 USA
G
lobal concern about securing water for multiple
uses has grown in recent years with quickly rising water
demands. With an increasing population, land conversion—and
urban development in particular—is inevitable in many parts of
the world. More than half of the world’s population lives in urban
areas, and it is expected that in the next 20 years, that figure will
rise to as much as 60% of the total population (UN Habitat,
2008). Therefore, an understanding of the interactions of landscape changes and water resources in the context of a rising population is imperative.
The southern United States represents a prime example of
these interactions and is expected to have the largest absolute
increase in population (an additional 43 million people) of any
region in the United States from 2000 to 2030 (U.S. Census
Bureau, 2005; Clouser, 2006). Recent water shortages in the
Southeast, which were exacerbated by the regional drought
from 2005 to 2007, have been linked with the growing regional
population (Center for Biological Diversity, 2009; Seager et
al., 2009). Urban expansion is expected to claim increasing
amounts of forested land in the Atlantic and Gulf Coasts and
the Southern Appalachian Piedmont in the coming years (Wear,
2002). For these reasons, the U.S. Southeast exemplifies the
common imbalance between water supply and demand as rising
population, land-use change, and climate interact to create periodic water shortages (Sun et al., 2008).
Water quantity and quality are important to meet water
demands for drinking, irrigation, recreational, and other uses.
The total areal coverage and spatial arrangement of forests in the
landscape influence water quantity and quality. There is a direct
link between forest and water resources, with 70 to 80% of flowing water in the United States originating as precipitation in forests (Binkley et al., 2004). Furthermore, historical and current
patterns of land use and cover within a watershed also have the
potential to critically influence water resources. Based on a review
of published studies, this paper discusses myriad factors that influence water resources, identifies which factors are most critical, and
develops general conclusions regarding water quality and quantity
in the Southeast.
R.C. Nagy, Dep. of Ecology and Evolutionary Biology, Brown Univ., 80 Waterman
St., Providence, RI 02912; B.G. Lockaby and L. Kalin, School of Forestry and Wildlife
Sciences, Auburn Univ., 602 Duncan Dr., Auburn, AL 36849; B. Helms, Biological
Sciences, Auburn Univ., 101 Life Sciences Building, Auburn, AL 36849; D. Stoeckel,
College of Veterinary Medicine, Auburn Univ., 105 Green Hall, Auburn, AL 36849.
Assigned to Associate Editor H.M. Selim.
Abbreviations: IS, impervious surface; SRP, soluble reactive phosphorus; TDS, total
dissolved solids; TSS, total suspended solids.
867
Characteristics of the U.S. Southeast
The climate of the Southeast is humid subtropical, with an average annual temperature range of 15 to 21°C and a precipitation
range of 101 to 152 cm yr−1 (Bailey, 1980). The region can be
divided broadly into three physiographic provinces: the Coastal
Plain, the Piedmont, and the southern Appalachians (Fig. 1).
As noted by Jackson et al. (2004), the lower Coastal Plain is
characterized by relatively high precipitation (>120 cm yr−1),
high evapotranspiration (>65% of precipitation), and a low
runoff ratio (<30%), or the runoff volume divided by the rainfall volume, due to the inherent flatness of much of the region.
The upper or hilly Coastal Plain and Piedmont may have less
precipitation (<120 cm yr−1) but has a higher runoff ratio (20–
60%) than that of the Coastal Plain as topography becomes
steeper. Finally, the southern Appalachians may exhibit higher
rainfall (as high as 200 cm yr−1) than the other two regions along
with low evapotranspiration and a higher runoff ratio (>50%)
compared with the Piedmont or Coastal Plain.
Coastal Plain relief is primarily level along much of the Atlantic
Coast and the lower Coastal Plain along the Gulf of Mexico.
However, the upper Coastal Plain of Alabama and Mississippi is
moderately to gently rolling (Martin and Boyce, 1993). Coastal
Plain soils are predominantly sandy in nature. The Piedmont is
characterized by gently rolling to steep terrain and clayey surface
and subsurface soils. As a result of the terrain and soil texture,
erosion potential is high throughout the Piedmont and increases
with increasing proximity to the Blue Ridge (Appalachians)
(Trimble, 2008). The Southern Appalachians reflect wide variation in elevation (225–900 m) and steep topography. Soil parent
material here is diverse and includes sandstone, shale, and a
variety of other materials. Consequently, soil particle size varies
widely as well (Martin and Boyce, 1993). Overall, the predominant soil order in the Southeast is Ultisols, which are strongly
leached and nutrient poor with a subsurface accumulation of
clay (Bailey, 1980; NRCS, 2009).
The natural climax vegetation of uplands in the Southeast
consists of forests of evergreen or live oak (Quercus virginiana,
Quercus geminata, and Quercus laurifolia), laurel (Laurus nobilis and Persea borbonia), magnolia (Magnolia grandiflora and
Magnolia virginiana), and sweetgum (Liquidambar styraciflua)
(Bailey, 1980). Swamps along the Atlantic Coast contain cypress
Fig. 1. Physiographic regions of the U.S. Southeast.
868
(Taxodium distichum var. distichum and Taxodium distichum
var. nutans) and gums (Liquidambar styraciflua and Nyssa sylvatica). Depressions in savannas, known as bogs or pocosins, are
dominated by evergreen shrubs (Bailey, 1980; Ainslie, 2002).
Flatwoods areas in the Coastal Plain are dominated by Pinus sp.
The Southeast has undergone extensive changes in land use,
with a notable period of agricultural expansion that began in the
19th century and ended by 1930 (Wear, 2002; Trimble, 2008).
A period of agricultural abandonment and subsequent reforestation ensued. Forest industry gained prominence throughout the region in the early 20th century (Wear, 2002). More
recently, agricultural and forested lands have been converted to
urban land uses (Table 1). In 2002, urbanization was identified
as the leading cause of forest loss within the region. Increases in
urban land use are expected for the Southeast in coming years to
accommodate the growing population (Wear, 2002).
Hydrological Effects
Watershed studies have elucidated the relationship between
forests and water resources. Reduced forest cover decreases
evapotranspiration, and thus increases in streamflow are
often observed after forest cutting (Hibbert, 1967; Bosch and
Hewlett, 1982). Forest harvesting increased streamflow from
69 to 210 mm yr−1 in a study that included 13 Southern states
(Grace, 2005). Conversely, reforestation can decrease water
yield. As an example, streamflow decreases of 76 to 152 mm
yr−1 were observed after 75% reforestation of the Pine Tree
Branch watershed in western Tennessee (Bosch and Hewlett,
1982). The precipitation regime influences the relative effects
after forest conversion, and absolute changes in streamflow are
greatest in areas with the highest precipitation (Hibbert, 1967).
The dominant tree species or functional type further influences
the change in water yield after forest harvesting or regrowth.
For example, conversion of hardwood to pine in the southern
Appalachians has been reported to decrease annual streamflow by as much as 20% and to dampen high- and low-flow
oscillations in the hydrograph (Jackson et al., 2004). Bosch
and Hewlett (1982) reported that, for every 10% decrease in
forest cover, an average 40-mm increase in water yield can be
expected for pine forests, and an average 25-mm increase can
be expected for deciduous hardwoods (Table 2).
Several studies have proposed a threshold of forest cover
removal, which, when exceeded, leads to a measurable difference in water yield (or another parameter). According to
Stednick (1996), 20% removal of forest cover altered water
yield in the Appalachians, but the threshold may be as high
as 45% in the Coastal Plain or Piedmont. However, Stednick
(1996) and Bosch and Hewlett (1982) noted that few studies
had examined changes in water yield under conditions of light
forest removal (<20%). It is possible that the lack of hydrologic
response data reflecting <20% removal may contribute to that
percentage’s identification as a threshold.
Temporal variation is an important consideration regarding
potential effects of changes in forest cover on water yield. The
observed pattern of decreased streamflow after afforestation is
not permanent. After a forest has reached maturity (∼30 yr for
pines), streamflow may approach the pre-afforestation level due
primarily to reduced transpiration rates and to a lesser extent
Journal of Environmental Quality • Volume 40 • May–June 2011
reduced leaf area index and therefore reduced interception (Scott
and Prinsloo, 2008). Results from Hubbard Brook, Marcell,
Fernow, and Leading Ridge Experimental Forests in the northeastern United States indicated that with natural regrowth, water
yield increases may diminish within 3 to 10 yr after forest cutting (de la Crétaz and Barten, 2007). This suggests that longer
rotations may be able to mitigate the negative effects of afforestation on water yield. Also, after deforestation, the effects on water
yield are promptly apparent, whereas the effects of reforestation
must be observed as the forest develops over time (Douglass and
Swank, 1972; Andréassian, 2004). Thus, studies lasting only a
few years may not capture the full hydrological effects of forest
establishment and growth. More long-term studies are needed
to quantify changes in water yield over time after afforestation.
Increases in streamflow and peak flows have been observed
after conversion of forest to agriculture (Hibbert, 1967),
whereas the effects of agricultural conversion on baseflow are
less clear. With agricultural conversion, soils may become compacted, thus reducing infiltration and increasing overland flow.
Increased overland flow can decrease groundwater recharge,
but evapotranspiration is also reduced, which can increase
recharge. Zhang and Schilling (2006) reported increased
streamflow (3–13 mm mo−1) and baseflow (4–12 mm mo−1)
with conversion of perennial vegetation to seasonal row crops
in the Mississippi River basin. For comparison, peak discharge
in tilled land was 12 times greater than peak discharge in a
forested pasture in southwestern Wisconsin (de la Crétaz and
Barten, 2007). As a consequence of increased flow rates in
agricultural lands, erosion is often increased and can be highly
problematic (discussed below).
Similar to the conversion of forest to agriculture, conversion
of forest to urban land uses greatly reduces infiltration due to
the introduction of impervious surfaces (IS), which can increase
runoff velocity (Table 3). Infiltration was reduced by 60% as
a result of rapid urbanization in the White Rock watershed in
Texas, which changed from 87% rural in 1961 to 95% urban
by 2002 (Vicars-Groening and Williams, 2007). As a result of
reduced infiltration and increased runoff velocity, urban flow
hydrographs may entail more rapid and larger pulses (i.e.,
becoming less stable or more flashy in nature) (e.g., Fig. 2); this
is a reflection of more frequent and intense flooding in urban
compared with forested areas (Beighley et al., 2003; Calhoun et
al., 2003; Schoonover et al., 2006; Crim, 2007; de la Crétaz and
Barten, 2007). In Houston, Texas, increases in watershed imper-
Table 1. Major land uses (as of 2002) by state of the U.S. Southeast.†
State
Alabama
Arkansas
Florida
Georgia
Kentucky
Louisiana
Mississippi
North Carolina
Oklahoma
South Carolina
Tennessee
Texas
Virginia
Cropland
Pasture, range
Forest
Land use
Transportation, industrial,
parks, wildlife
Urban
Other
———————————————————————————— % ————————————————————————————
11.51
5.87
70.58
4.68
3.51
3.85
28.94
6.94
55.14
4.60
1.75
2.63
10.77
13.62
42.41
13.06
11.47
8.67
13.61
3.31
64.23
5.38
6.49
6.97
32.85
9.06
46.99
6.32
3.08
1.70
19.15
6.14
49.22
6.91
3.84
14.75
20.20
7.40
61.86
3.19
2.00
5.35
16.68
2.81
59.87
7.68
7.36
5.61
34.65
39.79
14.18
3.72
1.68
5.97
12.07
2.33
63.83
5.57
6.23
9.97
26.86
4.53
51.77
8.98
5.94
1.93
24.14
58.65
7.03
3.27
2.74
4.19
16.45
5.42
60.66
6.27
6.02
5.18
† Data from http://www.ers.usda.gov/data/majorlanduses/.
Table 2. Examples of hydrological, physicochemical, and bacterial responses of water resources with changes in forest cover and introduction of
impervious surfaces.
Land use/cover change
10% change forest
10% change forest
>10% IS cover (vs. forested)
10–20% IS cover (vs. forested)
20–30% nonforest (vs. >90% forest)
20–30% nonforest (vs. >90% forest)
20–30% nonforest (vs. >90% forest)
20–30% nonforest (vs. >90% forest)
>13% IS cover (vs. forested)
>13% IS cover (vs. forested)
3–4% urban (vs. forested)
Response
40 mm change water yield—pines
25 mm change water yield–deciduous hardwoods
32% of 159% increase in peak flows
2× increase in runoff
3× increase in TSS†
2–3× increase in TDS
4–5× increase in turbidity
5–8× increase in NO3 concentration
3.8–7.3× increase in median NO3 concentration
3× increase in median Cl concentration
4–5× increase in fecal coliform concentration
in baseflow
Location
Source
many
Bosch and Hewlett (1982)
many
Bosch and Hewlett (1982)
Houston, Texas
Olivera and DeFee (2007)
many
Paul and Meyer (2001)
Blue Ridge physiographic region
Price and Leigh (2006)
Blue Ridge physiographic region
Price and Leigh (2006)
Blue Ridge physiographic region
Price and Leigh (2006)
Blue Ridge physiographic region
Price and Leigh (2006)
west Georgia
Crim (2007)
west Georgia
Crim (2007)
Appalachian Mountains,
Bolstad and Swank (1997)
North Carolina
† TDS, total dissolved solids; TSS, total suspended solids.
Nagy et al.: Water Resources and Land Use in a Humid Region
869
Table 3. Examples of changes in hydrologic flows expected with increasing imperviousness.†
Component
Evapotranspiration
Runoff
Shallow infiltration
Deep percolation
Total
Imperviousness
0%
10–20%
35–50%
40%
10%
25%
25%
100%
38%
20%
21%
21%
100%
35%
30%
20%
15%
100%
75–100%
30%
55%
10%
5%
100%
† Data from Paul and Meyer (2001).
vious cover contributed to a 159% increase in peak flows (Table
2), although a greater portion of this increase was attributable to
precipitation than increased watershed IS (Olivera and DeFee,
2007). Hydrographs from urban watersheds near Columbus,
Georgia and Birmingham, Alabama showed that during storm
events, discharges were flashier than those from forested watersheds (Schoonover et al., 2006; McMahon et al., 2003).
Fig. 2. Examples of (a) forested and (b) urban hydrographs from west
Georgia. Adapted from Crim (2007).
870
Urbanization also reduces the contribution of baseflow to
streamflow and decreases flows in the intervals between storms
(Rose and Peters, 2001; Wang et al., 2001; Calhoun et al.,
2003). For example, the contribution of baseflow to streamflow
in Peachtree Creek near metropolitan Atlanta, Georgia has been
reduced from 50 to 30% as the urban area has expanded in recent
decades (Calhoun et al., 2003). Rose and Peters (2001) reported
that low-flows in urban watersheds near Atlanta were 25 to 35%
lower than the low-flows in less urbanized watersheds. Thus,
in urban watersheds, the high-flow and the low-flow extremes
are enhanced (Lazaro, 1990; Moscrip and Montgomery, 1997;
Rose and Peters, 2001), and hydrographs reflect this instability. However, low-flows may be compensated somewhat by discharges from wastewater plants, and thus the effects on baseflow
can be somewhat variable (Paul and Meyer, 2001).
With conversion of forests to urban and agricultural uses,
water supply will become a more pressing issue in the Southeast.
Sun et al. (2008) used hydrological, demographic, climate
change, and land-use change models to illustrate the interactive
effects of these drivers on water stress through the year 2020.
In that study, water supply stress depended greatly on which
climate scenario was used, emphasizing the need for accurate,
detailed climate change predictions. The scenario simulating
population growth only suggested that increased water stress
is likely near Raleigh (North Carolina), Atlanta (Georgia), and
Tallahassee (Florida) due to fast growing populations—and
thus demand—in these cities. The land-use change-only scenario predicted slight decreases in water stress in Florida and
the Piedmont due to increased water yield after forest conversion to urban uses. However, the combined scenario with
climate change, population growth, and land-use change predicted increased water stress in the Southeast because climate
change and population growth had stronger effects on water
supply than land-use change. Although not accounted for in
that study, the authors point out that water availability will
also be affected by water quality, and thus these considerations
must be included in further model developments (Sun et al.,
2008). Updated projections of water stress are necessary as
climate change, land-use change, and population growth estimates are refined.
Although water yield may be reduced in forested watersheds
compared with urban watersheds, the percent of water available
for use may be greater in forested watersheds. Larger volumes
of water are transported through watersheds more rapidly in
urban areas (as indicated by rising limbs with steeper slopes on
the hydrograph), but access to that water for use may be limited.
For example, water available for consumption (Wc) was
modeled for two small watersheds in close proximity to
Birmingham, Alabama as a function of withdrawal capacity
Journal of Environmental Quality • Volume 40 • May–June 2011
(Ww) and the minimum flow required for environmental purposes (Qmin). The watersheds were similar in size but differed in
terms of imperviousness (Site 1 = 66% forest, 16% urban; Site
2 = 6% forest, 91% urban). The Qmin was set equal to Q10%, or
flow that was exceeded 90% of the time. An estimate of Ww
was derived from data for similar areas in Alabama and was
approximately 809 L d−1 person−1 or 0.47 m3 s−1 for a population of 50,000. Flow was standardized based on watershed
area. The Wc value was estimated over a 60-d period from 16
Aug. 2009 to 15 Oct. 2009, a period that was unusually wet
in Alabama. Proportions of Wc ranged from 24 to 60% in the
low IS watershed to between 17 and 38% for the more heavily
urbanized catchment. Differences between Site 1 (low urbanization) and Site 2 (the more urbanized watershed) were more
pronounced at higher withdrawal rates (Fig. 3).
Physicochemical Effects
Alterations in forest cover can change soil erosion and sediment deposition processes. Erosion often accompanies forest
conversion to other land uses because the stabilization of soils
provided by vegetation has been diminished. This is especially
the case when riparian areas are cleared. The extent of forest
loss within the watershed determines the degree of sediment
erosion. Price and Leigh (2006) compared streams from lightly
impacted (>90% forest) and moderately impacted (70–80%
forest) watersheds in the Blue Ridge physiographic region.
Total suspended solids (TSS), total dissolved solids (TDS), and
turbidity were higher (3×, 2–3×, and 4–5× for TSS, TDS, and
turbidity, respectively) in moderately impacted streams than
in lightly impacted streams (Table 2) (Price and Leigh, 2006).
Similarly, Sutherland et al. (2002) found that streams with 13
to 22% nonforested catchments (disturbed streams) had significantly higher turbidity compared with catchments with <3%
nonforested area (reference streams). This pattern was observed
in baseflow and in stormflow, although the difference between
disturbed and reference stream turbidity was slightly greater in
stormflow (Sutherland et al., 2002). Generally, there is a higher
percentage of smaller-sized particles with increasing nonforested area (Sponseller et al., 2001; Sutherland et al., 2002).
Watershed forest cover also greatly influences stream nutrients. Riparian vegetation and soils may act as a biogeochemical
or physical filter by assimilating nutrients in runoff before the
water reaches the stream. Nitrate (NO3−) concentrations were
higher in watersheds with greater nonforested area in the Little
Tennessee River (Price and Leigh, 2006), the Upper Tennessee
River (Scott et al., 2002), and Coweeta Creek (Bolstad and
Swank, 1997). Similarly, in the Upper Roanoke River Basin,
total inorganic nitrogen ranged from 0.105 to 0.932 mg L−1
and increased as the percentage of nonforested area in the
catchment increased (R2 = 0.59; p = 0.02) (Sponseller et al.,
2001). Ammonium (Price and Leigh, 2006), orthophosphates
(Price and Leigh, 2006), and soluble reactive phosphorus (SRP)
(Sponseller et al., 2001) did not show differences as forest cover
varied. In the study by Price and Leigh (2006), ammonium and
orthophosphate concentrations were below detection limits;
consequently, subtle changes could not have been detected.
Sponseller et al. (2001) noted that variation in water chemistry
Fig. 3. Water availability (%) in urban (Site 1) vs. forested (Site 2)
watersheds near Birmingham, Alabama with increasing withdrawal
rates and a minimum flow (Qmin) amount of 10%. Unpublished data
from L. Kalin.
was influenced more strongly by geomorphology than by land
use.
Conversion of forests for agricultural use has been linked to
erosion and subsequent sediment deposition in the Southeast,
particularly in the Piedmont. Agricultural land use during the
cotton era, ending around 1920, was extensive enough to erode
an average of 10 to 30 cm of topsoil from the entire Piedmont
physiographic region (Trimble, 2008). Historic land use in the
Piedmont reportedly contributed to 1.6 m of sediment deposition in the Murder Creek floodplain (Jackson et al., 2005) and
an average of 1.79 m of sediment deposition in Bonham Creek
and Sally Branch watersheds (Fig. 4) (Casarim, 2009). Similar
patterns have been reported in the Northeast, including a study
near New Haven, Connecticut, where sediment deposition
increased 40% after European settlement in New England (de
la Crétaz and Barten, 2007). Trimble (2008) noted that, as erosive land uses (particularly agriculture) declined in the southeastern Piedmont after the 1930s, stream turbidity declined
sharply. Sediment deposition can be greater in large river systems, such as the floodplains of the Roanoke River in North
Carolina, where up to 6 m of sediment deposition has been
recorded (Dr. Cliff Hupp, USGS, personal communication).
Fig. 4. Sediment deposition in west Georgia Piedmont streams. Photo
from F. Casarim.
Nagy et al.: Water Resources and Land Use in a Humid Region
871
Eroded sediment can also be deposited in stream channels and
gradually exported over time, complicating the attribution of
sediment erosion to any individual event. For example, differentiation of current versus historical land-use practices requires
isotope or other fingerprinting techniques (Walling, 2005).
Lastly, in agreement with the comparison of sediment particle
size in forested vs. nonforested area, as agricultural intensity
increased, the percentage of silt increased in the Upper Little
Tennessee River (Hagen et al., 2006).
Forest conversion to agricultural land may shift a landscape
from biogeochemical sink to source activity. Higher stream
nutrient concentrations are common in agricultural areas
due to fertilizer application and riparian vegetation removal.
Higher NO3− concentrations have been found in agricultural
watersheds than in forested watersheds in the Appalachians
(Hagen et al., 2006), the Piedmont (Crim, 2007), and the
Coastal Plain (Lowrance et al., 1985; Lehrter, 2006). For
example, NO3− concentrations increased fourfold along a gradient from forest to heavy agriculture (Hagen et al., 2006).
In Maine, mean NO3− concentrations were 0.10 and 2.0 mg
L−1 for forested and agricultural watersheds, respectively (de la
Crétaz and Barten, 2007). In the southeastern Coastal Plain,
Lowrance et al. (1985) found higher loads of chloride (Cl−) and
base cations such as potassium (K+), calcium (Ca2+), and magnesium (Mg2+) in watersheds with more agricultural use. In
contrast, Lowrance et al. (1985) and Crim (2007) found that
phosphorus was not affected by the proportion of crop and
pasture area within Coastal Plain and Piedmont watersheds,
respectively. However, mean SRP concentrations doubled
along the forest-to-agriculture gradient in the Appalachians
(Hagen et al., 2006).
Urban land use promotes increased sediment erosion compared with forested land due to the presence of IS and increased
runoff velocities (Lenat and Crawford, 1994; Paul and Meyer,
2001; Schoonover et al., 2005; Clinton and Vose, 2006). Total
suspended solids concentrations were five times higher in an
urban stream than in the reference stream in the Appalachians
(Clinton and Vose, 2006), and TDS concentrations were twice
as high in urban streams compared with forested streams in the
west Georgia Piedmont (Schoonover et al., 2005; Crim, 2007).
Sediment export may be more associated with stormflow than
baseflow, but this is not always the case. Additionally, construction practices can be particularly conducive to increased erosion, with large bursts of sediment movement occurring during
flood events. Paul and Meyer (2001) reported that sediment
yields from developing catchments during stages of construction can be two to four orders of magnitude larger than yields
from forested catchments. However, in watersheds with old or
stable urban land use, streams may become incised and sediment starved because higher flows have scoured and downcut
the stream channel over time or because sediment sources are
diminished when channels are covered in concrete (Groffman
et al., 2003). In contrast to the previously explained patterns
of smaller sediment particles with increasing nonforested area,
Holland et al. (2004) and Lerberg et al. (2000) found either
no significant relationship or fewer fine particles with increasing urban area in the Coastal Plain. The contrasting results
associated with particle size exportation may reflect differences
872
among the clayey Piedmont surface soils vs. the sandy counterparts of the Coastal Plain.
Sources of nutrients in urban areas include fertilizer and
wastewater discharge. Median NO3− concentrations in urban
areas (>24% IS) were significantly greater than in nonurban
areas (<5% IS) in baseflow (1.64 and 0.61 mg L−1 for urban
and nonurban areas, respectively) and stormflow (1.93 and
0.36 mg L−1 for urban and nonurban areas, respectively)
(Schoonover and Lockaby, 2006). Ammonium concentrations were significantly greater in urban streams than nonurban streams in stormflow only (0.15 mg L−1 for urban vs.
0.00 mg L−1 for nonurban) (Schoonover and Lockaby, 2006).
Streams may display nutrient responses at low levels of urban
development. Again the concept of a threshold arises, but here,
rather than a threshold to forest removal, it is a threshold to
increasing IS cover. Some researchers have suggested that 10%
imperviousness is the level at which effects on water resources
can begin to be observed and that once 30% imperviousness is
reached, the results can be severe (Fig. 5a) (Arnold and Gibbons,
1996; Calhoun et al., 2003). Yet, pronounced increases in sulfate
(SO42−), Na+, and K+ concentrations, pH, and a general chemical index were reported in Piedmont streams at <5% watershed
imperviousness (Harned and Cuffney, 2005; Crim, 2007).
These seeming contradictions regarding minimum thresholds
may be a function of how those are defined. For example, the
stylized response graph from Arnold and Gibbons (1996) (Fig.
5a) reflects a linear relationship between imperviousness and
stream condition, with thresholds identified based on subjective
criteria. Alternatively, some studies have shown the response
relationship to be curvilinear (Fig. 5b), with thresholds identified at points of inflection on the response curve. The latter
approach may tend to yield lower estimates.
Aquatic Biota
Land conversion greatly affects the physical stream environment. Decreased forest cover, particularly in riparian areas,
can increase water temperatures and subsequently decrease
dissolved oxygen, creating inhospitable living conditions for
aquatic organisms. Furthermore, sediment and nutrient inputs
from agricultural and urban lands can be excessive. Increased
nutrients can cause algal blooms or eutrophication and further
deplete dissolved oxygen. Also, altered hydrological conditions,
such as increased discharge, can greatly influence instream conditions. Together with the associated changes in habitat, all of
these effects place stress on aquatic organisms.
Some organisms are better suited to handle these stresses
than others; thus, changes in community assemblages frequently
accompany land-use and cover changes. Typical responses
include increased abundance of exotic and tolerant species,
reduction of sensitive species and overall diversity, and altered
productivity (Meyer et al., 2005; Walsh et al., 2005). Most of the
literature cites a range from 8 to 20% IS and 30 to 50% agriculture as thresholds for biota (Paul and Meyer, 2001; Allan, 2004).
These values vary with the response variable (macroinvertebrates,
fish, etc.) and with region studied. As such, variable and location issues impede the broad utility of thresholds for monitoring
and management (Groffman et al., 2006). There is also evidence
Journal of Environmental Quality • Volume 40 • May–June 2011
that these relationships are too complex for a single threshold to
apply to biotic parameters (Allan, 2004).
After forest conversion to urban and agricultural uses,
periphyton diversity may be reduced while biomass is increased
due to the increase of stream nutrients or temperature (Biggs,
1996; Burkholder, 1996; Sponseller et al., 2001). For example,
maximum temperature explained 73% of the variation in chlorophyll a in southwestern Virginia streams (Sponseller et al.,
2001). The relationship with diversity may be more complicated than originally thought. At moderate levels of nutrient
enrichment, algal species richness and diversity may increase
(Delong and Brusven, 1992; Chessman et al., 1999). However,
past a certain point of nutrient enrichment, diversity decreases,
and assemblages are often dominated by only one or several
taxa (Lobo et al., 1995; Juttner et al., 1996).
With land-use change, macroinvertebrates display similar
patterns to periphyton, including reduced diversity and richness (Pratt et al., 1981). There may also be a shift to more tolerant organisms (Lenat and Crawford, 1994). With extensive
land conversion, assemblages that act functionally as shredders may be replaced by collector or scraper functional feeding groups (Roy et al., 2003; Hagen et al., 2006; Helms et
al., 2009). However, in low- and moderate-intensity agriculture in the Upper Little Tennessee River watershed, shredder
abundance increased in response to altered habitat conditions
(increased light, higher water temperatures) and ample nutrients and food supplies (Hagen et al., 2006).
Land conversion and habitat alteration have been linked with
marked changes in mollusk populations in the Southeast. Direct
(channelization, dams, dredging) and indirect (land use in the
floodplain) changes to streams can degrade habitat for mollusk
populations (Gillies et al., 2003). Examples of anthropogenic
effects on mollusks include declines in mussel abundance in the
Line Creek watershed in Georgia (Gillies et al., 2003) and the
Coosa River watershed covering parts of Alabama, Georgia, and
Tennessee (Gangloff and Feminella, 2007). However, in watersheds where agricultural land use is widespread and somewhat
uniform, geologic and landscape properties, such as parent material, soil type, and topography, can be better predictors of mollusk populations than the current land use (Allan, 2004).
Similar to the other organisms discussed, fish are affected by
hydrological and morphological stream changes, which alter
habitat conditions. For example, increased siltation and sedimentation from forest conversion can homogenize habitats.
Greater siltation and sedimentation have also been linked with
decreased abundance of benthic feeding fish and endemic species (Berkman and Rabeni, 1987; Roy et al., 2005), increased
abundance of cosmopolitan species (Walters et al., 2003; Roy
et al., 2005), and reduced fish spawning success (Sutherland et
al., 2002). In west Georgia, Helms et al. (2005) found signs of
degraded fish health, including a positive correlation between
an increase in tumors and lesions (p < 0.001) and % urban area
and a negative correlation between biotic integrity (p < 0.001)
among fish species and % urban area. Lastly, nonnative fish
species can be introduced intentionally for recreation and commercial purposes or unintentionally by the displacement of fish
in other areas (Marsh and Minckley, 1982).
Due to their dual habitats in aquatic and terrestrial systems,
amphibians are highly sensitive to changes in land use and
Nagy et al.: Water Resources and Land Use in a Humid Region
Fig. 5. Comparison of physicochemical responses of streams to
increases in impervious surfaces. (a) Stylized linear response of stream
health (adapted from Arnold and Gibbons, 1996). (b) Curvilinear
response of sulfate concentration (adapted from Crim 2007).
cover, especially changes within the riparian zone. Amphibian
abundance may be a function of the width and the length of
the riparian buffer, although breaches in the stream buffer
may mask a relationship between abundance and buffer width
(Miller et al., 2007). In the Coastal Plain of Georgia, Muenz
et al. (2006) found increased instream larval salamander abundance in buffered streams but did not find a relationship for
total amphibian abundance, presence, or absence with buffered vs. unbuffered streams. In the North Carolina Piedmont,
southern two-lined salamander (Eurycea cirrigera) populations
have decreased by about 44%, and northern dusky salamander (Desmognathus fuscus) populations have decreased as much
as 30% in the last three decades with increasing urbanization
(Price et al., 2006). Hydrological changes that accompany land
conversions are also important for amphibians. Barrett and
Guyer (2008) reported that with increased flood frequency,
salamander density and survivorship decreased in the Georgia
Piedmont. Furthermore, increased spate flow velocities associated with urbanization can decrease larval retention but result
in a higher growth rate for those animals remaining (Barrett et
al., 2010a; Barrett et al., 2010b). Increased peak flows, reduced
base flows, and loss of rocky habitat were contributing factors
to declines in salamander abundance in urban streams in Wake
County, North Carolina (Miller et al., 2007).
Human and Environmental Health
Potential threats to human and environmental health are associated with the introduction of various contaminants from urban
873
and agricultural land uses, including sediment, pathogens, and
nutrients. The primary influence on future water quality in the
southern United States will be nonpoint-source pollution from
urban sprawl (West, 2002) and increased levels of pathogens,
heavy metals, pharmaceuticals, and other substances have been
shown to increase as watersheds become urbanized.
In urban and industrial streams in the South Carolina
Coastal Plain, sediment trace metal concentrations (Cu, Cr,
Pb, Zn, Cd, and Hg) were 2 to 10 times higher than in forested
streams (Sanger et al., 1999a). Metals are problematic because
they do not degrade over time but rather tend to accumulate in
the sediments of water bodies and floodplains (Klapproth and
Johnson, 2000). For example, studies on the Chickahominy
River in Virginia found Pb, Cr, Cu, Zn, Cd, Zn, and Sn in
floodplain sediments downstream of Richmond (Klapproth
and Johnson, 2000). Additionally, organic contaminants,
including polycyclic aromatic hydrocarbons, polychlorinated
biphenyls, and dichlorodiphenyltrichloroethane, were significantly higher in urban/industrial and some suburban streams
compared with forested streams (Sanger et al., 1999b). Lastly,
cumulative contaminant loading had a positive relationship
with % IS in the South Carolina Coastal Plain (R2 = 0.59; p <
0.0001) (Holland et al., 2004).
Another group of chemicals linked with increasing urban
land uses includes personal care products. Among these are perfumes, deodorants, pharmaceuticals, and hospital and veterinary
wastes. The widespread prevalence of these items has increased
drastically in the last few decades, and water filtration systems
and water regulations have not responded adequately to address
them (Kolpin et al., 2002; Bolong et al., 2009; Yu and Chu,
2009). For instance, ibuprofen (113 ng L−1) and an antimicrobial agent, triclosan (193 ng L−1), were detected in stream samples below a wastewater treatment plant on the West Prong Little
Pigeon River in eastern Tennessee (Yu and Chu, 2009). In a study
examining levels of antibiotics, prescription and nonprescription
drugs, steroids, and hormones in streams in Boston, New York,
Philadelphia, and Chicago, 80% of the streams had one or more
of the contaminants (de la Crétaz and Barten, 2007).
The results of the Clean Water Act included decreases in
municipal wastewater discharges (West, 2002), yet bacteria
counts remain high in urban areas. Specifically, populations
of Enterococcus, Escherichia coli, fecal coliform, and fecal streptococcus are often higher in urban streams than in forested
streams (Mallin et al., 2000; Schoonover et al., 2005; Clinton
and Vose, 2006; DiDonato et al., 2009). For example, as the
IS cover increased from about 5 to 25% in southeastern North
Carolina, fecal coliform density increased from 10 to 100 CFU
100 mL−1 (Mallin et al., 2000). Escherichia coli concentration was
also strongly correlated with IS cover (r = 0.973; p = 0.005) and
watershed population (r = 0.986; p = 0.002) (Mallin et al., 2000).
In a study near Columbus, Georgia, median fecal coliform concentrations were roughly an order of magnitude higher in urban
streams than in forested streams (1500 vs. 112 MPN 100 mL−1
for urban and forest, respectively, in baseflow and 2750 vs. 261.5
MPN 100 mL−1 for urban and forest, respectively, in stormflow)
and were significantly related to discharge (R2 = 0.15; p = 0.01
for urban streams) (Crim, 2007). Factors contributing to higher
bacterial concentrations in urban areas include elevated water
temperatures and increased turbidity as bacteria survival increases
874
with the opportunity to bind to sediment particles (Mallin et al.,
2000). In addition, designs of sewer systems may play a large role,
as evidenced in Columbus, Georgia, where combined stormwater–sewer overflows were associated with high fecal coliform
counts in some cases (Crim, 2007). Domestic animals may be a
contributing factor as well (Mallin et al., 2000).
Excess nutrients can greatly degrade water resources when
algal blooms are stimulated and dissolved oxygen is depleted.
Unpleasant odors, loss of recreational value, and toxicity to
fish and other aquatic organisms are results of eutrophication
caused by excess nutrients in water (Baird, 1990; Klapproth
and Johnson, 2000). These conditions have been reported
in the Pamlico, Neuse, and Chowan Rivers in eastern North
Carolina (Baird, 1990).
Pesticide use is prevalent throughout agricultural lands, and
herbicides account for 70% of the pesticides used today (de la
Crétaz and Barten, 2007). In the headwaters of the Suwanee River
Basin near the Georgia–Florida border, prometon, metalachlor,
and atrazine were detected in concentrations of 0.14, 0.09, and
0.07 μg L−1, respectively, and were elevated in the spring (Oaksford,
1994). However, most concentrations were below levels considered to be a health risk. Row crop pesticide application may
be fairly frequent and corresponds to crop cycles with preplant,
preemergent, and possibly postemergent application (Stell et al.,
1995). Pesticide application in the Apalachicola-ChattahoocheeFlint River Basin in Georgia may occur year-round, although rates
are highest from March to September (Stell et al., 1995). Pesticides
are commonly found in surface waters, and concentrations may
exceed the USEPA’s health standards, as was found for atrazine
concentrations in 27% of samples from the Mississippi, Ohio,
and Missouri Rivers (de la Crétaz and Barten, 2007). Although
atrazine is thought to be only slightly toxic, the long-term effects
are unknown (de la Crétaz and Barten, 2007). Additionally, the
potential of pesticides reaching groundwater resources is a serious
concern due to the difficulty of improving these resources once
they have been contaminated.
Physiographic Differences
Although general trends have been documented for the conversion of forest to agriculture and urban land uses (Fig. 6), the
influence of physiographic regions can be observed in hydrologic, sediment export, and aquatic biota responses. However,
for water chemistry, the effect of land use is stronger than that
of physiography in the Southeast. Topography and soil conditions are very influential regarding hydrology and sediment,
and land-use history also plays a role. The number of studies
conducted, and thus information available, is limited in the
Coastal Plains, whereas more research has been done in the
Appalachians and even more in the Piedmont. The effects of
land conversion on aquatic resources are more pronounced
in the Appalachians and Piedmont than in the Coastal Plain,
although only a few studies have done direct comparisons. For
example, greater sensitivity to land-use changes was reported
for streamflow in the Appalachians than in the Piedmont or
Coastal Plain (Stednick, 1996). Morgan and Cushman (2005)
and Utz et al. (2009) found stronger responses to land-use
change in fish and macroinvertebrates, respectively, in the
Piedmont than in the Coastal Plain.
Journal of Environmental Quality • Volume 40 • May–June 2011
Fig. 6. Conceptual model of the effects of land-use change on properties of receiving streams in (a) forest to agriculture and (b) forest to urban
(including historical changes). Solid arrows are contemporary drivers; dashed lines are historical drivers; pluses, minuses, and zeroes refer to
increases, decreases, and no effect, respectively, of each parameter (comma separators are sequential effects; slashes are either/or). Question
marks are unknowns. DO, dissolved oxygen; DOC, dissolved organic carbon; TDS, total dissolved solids; TSS, total suspended solids.
Nagy et al.: Water Resources and Land Use in a Humid Region
875
Conclusions
The “urban stream syndrome” is characterized by changes in
hydrology, stream chemistry, channel morphology, biotic richness, and community assemblages (Meyer et al., 2005; Walsh
et al., 2005). In general, urban streams have more flashy
hydrographs, increased nutrients and contaminants, incised
channels, reduced species richness, and increased prevalence of
tolerant organisms compared with forested streams. Even in
humid areas, changes in hydrology and water quality may have
significant negative effects on water supplies for growing populations, especially when rising demand coincides with periodic
drought. Although removal of forest cover usually results in
increased stream velocity and discharge, there may be problems
in urban areas with “capturing” sufficient water for human use.
Changes in hydrology, sediment export, and aquatic biota seem
to vary in terms of magnitude among physiographic regions as
follows: Appalachians > Piedmont > Coastal Plain. However,
for water chemistry, the influence of land use dominates that
of physiography in the Southeast.
The threshold of disturbance, as indicated by parameters
such as the % IS within a watershed, can be a useful guide to
estimate when effects on aquatic resources may be observed.
Previous studies have reported that stream degradation can
occur at 10 to 20% impervious cover, yet changes have been
observed at <5% IS in the Piedmont (Harned and Cuffney,
2005; Crim, 2007) and Southern Appalachians (Bolstad and
Swank, 1997; Clinton and Vose, 2006). Thus, although imperviousness may be a useful parameter to indicate the extent
of disturbance, the effects can be observed even at very low
levels of impervious cover in the early stages of development.
Furthermore, threshold variability seems to be dependent on
the ecoregion and on the response variable, complicating the
use of a single variable (e.g., % IS) as an estimate of disturbance. More recently, some researchers have used the termed
“suburban stream syndrome” to identify similar effects at lower
levels of IS cover and at the edges of urban areas (Cunningham
et al., 2009). For example, streams in Duchess County, New
York illustrated logarithmic increases in NO3− with increasing
IS cover (Cunningham et al., 2009).
Temporal Considerations
Historical land use can have an equal or greater influence than
current land use on the state of water resources. For example,
watershed land-use data from the 1940s and 1950s were better
predictors than contemporary land-use data in describing current stream biotic conditions (de la Crétaz and Barten, 2007;
Maloney et al., 2008). Scott et al. (2002) reported that historical land use had a greater effect on erosion and stream sediment, whereas current land use influenced stream nutrients and
temperature. Thus, it is important to be cognizant of historical
practices, such as heavy legacy sediment accumulation within
floodplains and stream channels, when evaluating current and
predicting future conditions of water resources.
Several studies, such as Schoonover and Lockaby (2006)
and Crim (2007), have indicated that water quality may be
as impaired under active and recent development compared
with areas that have been developed for much longer periods
of time. Consequently, the onset of water quality impairment
876
is relatively rapid once development begins, and, because there
is no vegetation recovery period as there is in the case of forest
harvesting, sediment, biogeochemical, and other types of terrestrial exports may remain high indefinitely. The ramifications
of development processes, including vegetation clearing and
construction, on water resources should not be underestimated.
Long-term studies are needed to capture temporal dynamics. Every land-use transition has a temporal effect, and thus
the response variables are likely to change over time. Future
studies could quantify how biotic populations change over
time after urban development and agriculture or conversely
after afforestation.
Spatial Considerations and Interactions
In addition to the total amount of IS, forests, and agricultural
land within the watershed, the spatial arrangement of these
lands determines the water resource response. Arnold and
Gibbons (1996) emphasized the need for assessment of the spatial arrangement of impervious surfaces, yet few studies seem
to incorporate this component. A combination of modeling
and field studies to examine different arrangements of forest,
urban, and agricultural land within a watershed is necessary to
inform land management decisions. Furthermore, within each
watershed or subwatershed, there are often multiple land uses
that interact to influence water resources. We need to understand the cumulative effects of multiple land uses on water
resources at various temporal stages of land conversion.
A common best management practice is maintaining or
restoring the riparian zone. Riparian forests create a buffer
between the terrestrial and aquatic systems and have many
associated functions in this transitional zone. Functional riparian zones with high hydrologic connectivity between streams
and uplands provide essential biogeochemical processing, plant
uptake, and physical entrainment of sediments that ultimately
lessen the effect of watershed disturbance (Naiman et al.,
2005). The filtering capability is due to their position in the
landscape and to geomorphic, hydrologic, and biotic processes
within riparian areas (Naiman et al., 2005). Additionally, even
in cases where riparian zones are left intact in urban areas, their
functioning ability may be inhibited (Groffman et al., 2003;
Walsh et al., 2005). For example, Walsh et al. (2005) explained
that channel incision, increased IS cover, and piped drainage
can lower the water table and thereby eliminate interactions
between the riparia and shallow groundwater, thus limiting the
capacity for filtration of pollutants. With regard to riparian restoration, the potential for nutrient and pollutant filtration and
flood regulation afforded by riparian forests must be weighed
against the short-term reduction in water yield that can occur
after reforestation.
The complex relationship between forests and water
resources requires attention from scientists and the general
public alike. Forests influence water quantity and quality, and
the relationships vary based on land-use history, geology, climate, and topography. Therefore, caution must be used in land
conversions that alter the location or extent of forests within
the watershed to minimize the effects on water quantity and
quality. Consideration of water and land-use interactions is
required to sustain water resource demands of rapidly growing
populations in humid and arid regions in coming decades.
Journal of Environmental Quality • Volume 40 • May–June 2011
Acknowledgments
Funding for this research was provided by the U.S. Forest Service
Southern Region, Coweeta Hydrologic Laboratory (Southern
Research Station), Southern Group of State Foresters, and the Center
for Forest Sustainability at Auburn University.
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Journal of Environmental Quality • Volume 40 • May–June 2011