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Streams in the Urban Landscape

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Abstract

The world’s population is concentrated in urban areas. This change in demography has brought landscape transformations that have a number of documented effects on stream ecosystems. The most consistent and pervasive effect is an increase in impervious surface cover within urban catchments, which alters the hydrology and geomorphology of streams. This results in predictable changes in stream habitat. In addition to imperviousness, runoff from urbanized surfaces as well as municipal and industrial discharges result in increased loading of nutrients, metals, pesticides, and other contaminants to streams. These changes result in consistent declines in the richness of algal, invertebrate, and fish communities in urban streams. Although understudied in urban streams, ecosystem processes are also affected by urbanization. Urban streams represent opportunities for ecologists interested in studying disturbance and contributing to more effective landscape management.
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Annu. Rev. Ecol. Syst. 2001. 32:333–65
Copyright c
°2001 by Annual Reviews. All rights reserved
STREAMS IN THE URBAN LANDSCAPE
Michael J. Paul1and Judy L. Meyer
Institute of Ecology, University of Georgia, Athens, Georgia 30602;
e-mail: mike@sparc.ecology.uga.edu, meyer@sparc.ecology.uga.edu
Key Words impervious surface cover, hydrology, fluvial geomorphology,
contaminants, biological assessment
Abstract The world’s population is concentrated in urban areas. This change in
demography has brought landscape transformations that have a number of documented
effects on stream ecosystems. The most consistent and pervasive effect is an increase
in impervious surface cover within urban catchments, which alters the hydrology and
geomorphology of streams. This results in predictable changes in stream habitat. In
additiontoimperviousness,runoff from urbanized surfaces aswell asmunicipal andin-
dustrialdischarges result in increased loading of nutrients, metals, pesticides, and other
contaminants to streams. These changes result in consistent declines in the richness
of algal, invertebrate, and fish communities in urban streams. Although understud-
ied in urban streams, ecosystem processes are also affected by urbanization. Urban
streams represent opportunities for ecologists interested in studying disturbance and
contributing to more effective landscape management.
INTRODUCTION
Urbanization is a pervasive and rapidly growing form of land use change. More
than 75% of the U. S. population lives in urban areas, and it is expected that more
than 60% of the world’s population will live in urban areas by the year 2030, much
of this growth occurring in developing nations (UN Population Division 1997, US
Census Bureau 2001). Whereas the overall land area covered by urban growth
remains small (2% of earth’s land surface), its ecological footprint can be large
(Folke et al. 1997). For example, it is estimated that urban centers produce more
than 78% of global greenhouse gases (Grimm et al. 2000) and that some cities
in the Baltic region claim ecosystem support areas 500 to 1000 times their size
(Boland & Hanhammer 1999).
This extensive and ever-increasing urbanization represents a threat to stream
ecosystems. Over 130,000 km of streams and rivers in the United States are im-
1Presentaddress: TetraTech,Inc.,10045Red Run Blvd., Suite110, Owings Mills,Maryland
21117.
0066-4162/01/1215-0333$14.00 333
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334 PAUL ¥MEYER
paired by urbanization (USEPA 2000). This makes urbanization second only to
agriculture as the major cause of stream impairment, even though the total area
covered by urban land in the United States is minor in comparison to agricul-
tural area. Urbanization has had similarly devastating effects on stream quality in
Europe (House et al. 1993).
Despite the dramatic threat urbanization poses to stream ecosystems, there has
not been a thorough synthesis of the ecological effects of urbanization on streams.
There are reviews discussing the impacts of a few aspects of urbanization [biology
of pollution (Hynes 1960), physical factors associated with drainage (Butler &
Davies2000), urban streammanagement (Baer & Pringle2000)] and a fewgeneral
reviews aimed at engineers and invertebrate biologists (House et al. 1993, Ellis &
Marsalek 1996, Suren 2000), but the ecological effects of urban growth on stream
ecosystems have received less attention (Duda et al. 1982, Porcella & Sorenson
1980).
An absolute definition of urban is elusive. Webster’s New Collegiate Dictionary
defines urban as “of, relating to, characteristic of, or constituting a city,” where the
definition of city is anything greater than a village or town. In human population
terms, the U. S. Census Bureau defines urban as “comprising all territory, popula-
tion, and housing units in urbanized areas and in places of 2,500 or more persons
outside urbanized areas,” where urbanized areas are defined as places with at least
50,000 people and a periurban or suburban fringe with at least 600 people per
square mile. The field of urban studies, within sociology, has a variety of def-
initions, which all include elements of concentrated populations, living in large
settlements and involving some specialization of labor, alteration of family struc-
ture, and change in political attitudes (Danielson & Keles 1985). In this review, we
relyon the census-baseddefinition, asitincludes suburbanareassurrounding cities,
which are an integral part of many urban ecological studies and represent, in many
cases, areas that will develop into more densely populated centers. However, many
industrial/commercial/transportation areas that are integral parts of urban and ur-
banizing areas have low resident population densities, but are certainly contained
within our view of urban areas.
Ecological studies of urban ecosystems are growing (McDonnell & Pickett
1990, USGS 1999, Grimm et al. 2000). A valuable distinction has been drawn
between ecology in cities versus ecology of cities (Grimm et al. 2000). The
former refers to the application of ecological techniques to study ecological sys-
tems within cities, whereas the latter explores the interaction of human and eco-
logical systems as a single ecosystem. Although our review focuses on stream
ecology in cities, it is our hope that it will provide information of value to the
development of an ecology of cities. The goal of this review is to provide a
synthesis of the diverse array of studies from many different fields related to
the ecology of urban streams, to stimulate incorporation of urban streams in
ecological studies, and to explore ecological findings relevant to future policy
development. This review is a companion to the review of terrestrial urban
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ECOLOGY OF URBAN STREAMS 335
ecosystems by Pickett et al. (2001). The review is structured in three parts that fo-
cus on the physical, chemical, and biological/ecological effects of urbanization on
streams.
PHYSICAL EFFECTS OF URBANIZATION
Hydrology
A dominant feature of urbanization is a decrease in the perviousness of the catch-
ment to precipitation, leading to a decrease in infiltration and an increase in sur-
facerunoff(Dunne & Leopold 1978). Asthe percent catchment impervious surface
cover (ISC) increases to 10–20%, runoff increases twofold; 35–50% ISC increases
runoff threefold; and 75–100% ISC increases surface runoff more than fivefold
over forested catchments (Figure 1) (Arnold & Gibbons 1996). Imperviousness
has become an accurate predictor of urbanization and urban impacts on streams
(McMahon & Cuffney 2000), and many thresholds of degradation in streams are
associated with an ISC of 10–20% (Table 1) [hydrologic and geomorphic (Booth
& Jackson 1997), biological (Klein 1979, Yoder et al. 1999)].
Various characteristics of stream hydrography are altered by a change in ISC.
Lag time, the time difference between the center of precipitation volume to the
center of runoff volume, is shortened in urban catchments, resulting in floods that
peak more rapidly (Espey et al. 1965, Hirsch et al. 1990). Decreases in flood
peak widths from 28–38% over forested catchments are also observed, mean-
ing floods are of shorter duration (Seaburn 1969). However, peak discharges are
higher in urban catchments (Leopold 1968). Flood discharges increase in pro-
portion to ISC and were at least 250% higher in urban catchments than forested
catchmentsin Texasand NewYorkaftersimilar storms (Espeyet al. 1965, Seaburn
1969). Flood discharges with long-term recurrence intervals are less affected by
urbanization than more frequent floods, primarily because elevated soil mois-
ture associated with large storms results in greater surface runoff in forested
catchments (Espey et al. 1965, Hirsch et al. 1990). Some exceptions to these
observations have been noticed, largely depending on the location of urbaniza-
tion within a catchment. If the ISC occurs lower in a catchment, flooding from
that portion can drain faster than stormflow from forested areas higher in the
catchment, leading to lower overall peak flood discharge and increased flood
duration (Hirsch et al. 1990). In addition, blocked culverts and drains, swales,
etc. may also detain water and lower peak flood discharges (Hirsch et al.
1990).
Afurther result ofincreased runoffisa reductionin theunit water yield:a greater
proportion of precipitation leaves urban catchments as surface runoff (Figure 1)
(Espey et al. 1965, Seaburn 1969). This reduces groundwater recharge and re-
sults in a reduction of baseflow discharge in urban streams (Klein 1979, Barringer
et al. 1994). However, this phenomenon has been less intensively studied than
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Figure 1 Changes in hydrologic flows with increasing impervious surface cover in
urbanizing catchments (after Arnold & Gibbons 1996).
flooding, and the effects of irrigation, septic drainage, and interbasin transfers
may mitigate the effects of reduced groundwater recharge on baseflow (Hirsch
et al. 1990). Baseflow may also be augmented by wastewater treatment plant
(WWTP) effluent. The Acheres (Seine Aval) treatment plant, which serves
8.1 million people, discharges 75 km west of Paris and releases 25,000 liters/s
during low flow periods (Horowitz et al. 1999), increasing baseflow discharge
in the Seine by up to 40% during low flow periods. More strikingly, wastewater
effluent constitutes 69% annually and at times 100% of discharge in the South
Platte River below Denver, Colorado (Dennehy et al. 1998). In our experience,
high percentage contributions of wastewater discharge to urban rivers are not
uncommon.
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ECOLOGY OF URBAN STREAMS 337
TABLE 1 Effects of impervious surface cover (ISC) resulting from urbanization on various
physical and biological stream variablesa
Study subject Findings Reference
Physical responses: hydrology
Streams in Texas Peak discharge increases and Espey et al. 1965
lag time decreases with ISC.
Streams in Pennsylvania Bankfull discharge increases Leopold 1968
and lag time decreases with
catchment ISC.
Review Surface runoff increases and Arnold & Gibbons
lag time decreases with 1996
increasing ISC (see Figure 1).
Streams in Washington Increase in bankfull discharge Booth & Jackson
with increasing ISC. At 10%, 1997
2 y urban flood equals a 10 y
forested flood.
Physical responses: geomorphology
Streams in Pennsylvania Channel enlargement increases Hammer 1972
with increasing ISC.
Streams in New York Channel enlargement begins Morisawa & LaFlure
at 2% ISC. 1979
Streams in New Mexico Dramatic changes in channel Dunne & Leopold
dimensions at 4% ISC 1978
Streams in Washington Channels begin widening at 6% Booth & Jackson
ISC; channels universally 1997
unstable above 10% ISC
Physical responses: temperature
Streams in Washington, DC Stream temperatures increase Galli 1991
with increasing ISC.
Biological responses: fish
Streams in Maryland Fish diversity decreased Klein 1979
dramatically above 12–15%
ISC and fish were absent
above 30–50% ISC.
Streams in Ontario, Fish IBI decreased sharply Steedman 1988
Canada above 10% ISC, but streams
with high riparian forest cover
were less affected.
Streams in New York Resident and anadromous fish Limburg & Schmidt
eggs and larvae densities 1990
decreased to 10% urban land
use and then were essentially
absent. (Continued)
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TABLE 1 (Continued)
Study subject Findings Reference
Streams in Maryland Fish diversity decreased Schueler & Galli
dramatically above 1992
10–12% ISC.
Streams in Wisconsin Fish IBI decreased rapidly Wang et al. 1997
at 10% ISC.
Streams in Ohio Fish IBI decreased rapidly Yoder et al. 1999
between 8% and 33% urban
land use.
Biological responses: invertebrates
Streams in Maryland Invertebrate diversity decreased Klein 1979
sharply from 1% to 17% ISC.
Streams in Northern Insect diversity decreased Jones & Clark
Virginia between 15% and 25% ISC. 1987
Streams in Maryland Insect diversity metrics moved Schueler & Galli
from good to poor at 15% ISC. 1992
Streams in Washington Insect IBI decreased sharply Horner et al. 1997
between 1% and 6% ISC,
except where streams had
intact riparian zones.
Streams in Ohio Insect diversity, biotic integrity Yoder et al. 1999
decreased between 8% and 33% ISC.
aIBI, index of biotic integrity.
Geomorphology
Themajorimpact ofurbanizationon basinmorphometryis analterationof drainage
density, which is a measure of stream length per catchment area (km/km2). Natural
channel densities decrease dramatically in urban catchments as small streams are
filled in, paved over, or placed in culverts (Dunne & Leopold 1978, Hirsch et al.
1990, Meyer & Wallace 2001). However, artificial channels (including road cul-
verts) may actually increase overall drainage densities, leading to greater internal
links or nodes that contribute to increased flood velocity (Graf 1977, Meyer &
Wallace 2001).
A dominant paradigm in fluvial geomorphology holds that streams adjust their
channel dimensions (width and depth) in response to long-term changes in sed-
iment supply and bankfull discharge (recurrence interval average=1.5 years)
(Dunne & Leopold 1978, Roberts 1989). Urbanization affects both sediment sup-
ply and bankfull discharge. During the construction phase erosion of exposed soils
increases catchment sediment yields by 102–104over forested catchments and can
be more exaggerated in steeply sloped catchments (Wolman 1967, Leopold 1968,
Fusillo et al. 1977). Most of this export occurs during a few large, episodic floods
(Wolman 1967). This increased sediment supply leads to an aggradation phase
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ECOLOGY OF URBAN STREAMS 339
Figure 2 Channel changes associated with urbanization. During the construction phase
of urbanization, hillslope erosion increases sediment supply leading to bed aggradation and
overbank deposition. After construction ceases hillslope sediment supply is reduced, but
bankfull flows are increased owing to increases in imperviousness. This leads to increased
channel erosion as channel incision and widening occur to accommodate increased bankfull
discharge.
as sediments fill urban channels (Figure 2). During this phase stream depths may
decrease as sediment fills the channel, and the decreased channel capacity leads to
greater flooding and overbank sediment deposition, raising bank heights (Wolman
1967). Therefore, overall channel cross-sections stay the same or even decrease
slightly (Robinson 1976). Ironically, the flooding associated with aggradation may
help attenuate increased flows resulting from increased imperviousness by stor-
ing water in the floodplain, temporarily mitigating urban effects on hydrography
(Hirsch et al. 1990).
After the aggradation phase sediment supply is reduced and geomorphic re-
adjustment initiates a second, erosional phase (Figure 2). High ISC associated
withurbanization increasesthe frequencyofbankfull floods,frequently byan order
of magnitude or, conversely, increases the volume of the bankfull flood (Leopold
1973, Dunne & Leopold 1978, Arnold et al. 1982, Booth & Jackson 1997). As
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a result, increased flows begin eroding the channel and a general deepening and
widening of the channel (channel incision) occurs to accommodate the increased
bankfull discharge (Hammer 1972, Douglas 1974, Roberts 1989, Booth 1990).
Increased channel water velocities exceed minimum entrainment velocities for
transporting bed materials, and readily moveable sediment is lost first as channels
generally deepen (Leopold 1973, Morisawa & LaFlure 1979). Channels may ac-
tually narrow during this phase as entrained sediment from incision is deposited
laterally in the channel (Dunne & Leopold 1978). After incision channels begin
to migrate laterally, bank erosion begins, which leads to general channel widening
(Booth 1990, Booth & Jackson 1997, Trimble 1997).
Duringthe erosional phasechannel enlargementcanoccur graduallyif increases
inwidth anddepth keeppace withincreases indischarge associatedwith increasing
ISC. In this case the channel enlargement may be barely noticeable (Booth 1990).
However,erosion more commonly occurs disproportionatelyto dischargechanges,
oftenleadingto bank failureandcatastrophicerosion inurbanstreams(Neller 1988,
Booth 1990). In developed urban catchments, as a result of this erosional readjust-
ment phase, the majority of sediment leaving the catchment comes from within-
channel erosion as opposed to hillslope erosion (Trimble 1997). The magnitude of
this generalized geomorphic response will vary longitudinally along a stream net-
work as well as with the age of development, catchment slope, geology, sediment
characteristics, type of urbanization, and land use history (Gregory et al. 1992).
Urban streams differ in other geomorphic characteristics from forested catch-
mentsas well.The spacingbetween pool-rifflesequences (distance between riffles)
is generally constant at 5–7 times channel width in forested catchments (Gregory
et al. 1994). Generally, this ratio stays constant in urban channels as they widen,
which means the absolute distance between pool-riffle units increases, although
there is some evidence that this spacing may decrease to 3–5 times channel width
(Gregory et al. 1994).
Changes in sediment supply may also alter channel pattern. Increased sediment
supply during construction has converted some meandering streams to braided
patterns or to straighter, more channelized patterns (Arnold et al. 1982). In the
latter case, channelizing leads to increased slope and therefore higher in-stream
velocities, especially where artificial channel alteration is carried out to increase
the efficiency of the channel in transporting flows (Pizzuto et al. 2000).
Urbanization can also alter sediment texture. Less fine sediment, increased
coarse sand fractions, and decreased gravel classes have been observed in ur-
ban channels as a result of alteration of sediment supply and altered velocities
(Finkenbine et al. 2000, Pizzuto et al. 2000). In addition to sediment changes,
large woody debris is also reduced in urban channels. Catchments in Vancouver,
British Columbia with greater than 20% ISC generally have very little large woody
debris, a structural element important in both the geomorphology and ecology of
Pacific Northwest stream ecosystems (Finkenbine et al. 2000).
Other geomorphic changes of note in urban channels include erosion around
bridges, which are generally more abundant as a result of increased road densities
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ECOLOGY OF URBAN STREAMS 341
in urban channels (Douglas 1974). Bridges have both upstream and downstream
effects,including plungepools created belowbridge culvertsthat may serveas bar-
riers to fish movement. Knickpoints are another common feature of urban chan-
nels. These readily erodeable points of sudden change in depth are created by
channel erosion, dredging, or bridge construction and are transmitted throughout
the catchment, causing channel destabilization (Neller 1988). Other features in-
clude increased tree collapse, hanging tributary junctions as a result of variable
incision rates, and erosion around artificial structures (e.g., utility support pilings)
(Roberts 1989).
Changes in the hydrology and geomorphology of streams likely affect the hy-
draulic environment of streams, altering, among other things, the velocity profiles
and hyporheic/parafluvial dynamics of channels. Such changes would affect many
ecologicalprocesses, fromfilter-feeding organisms(Hart &Finelli 1999)to carbon
processing and nutrient cycling (Jones & Mulholland 2000).
Temperature
Stream temperature is an important variable affecting many stream processes such
as leaf decomposition (Webster & Benfield 1986) and invertebrate life history
(Sweeney 1984). Urbanization affects many elements of importance to stream
heat budgets. Removal of riparian vegetation, decreased groundwater recharge,
and the “heat island” effect associated with urbanization, covered more fully in a
companion review (Pickett et al. 2001), all affect stream temperature (Pluhowski
1970), yet very little published data exists on temperature responses of streams
to urbanization. In one study on Long Island urban streams had mean summer
temperatures 5–8C warmer and winter temperatures 1.5–3C cooler than forested
streams. Seasonal diurnal fluctuations were also greater in urban streams, and
summertime storms resulted in increased temperature pulses 10–15C warmer
thanforested streams,a result of runoff fromheated impervioussurface(Pluhowski
1970). Similar effects on summer temperatures and daily fluctuations have also
been observed elsewhere (Table 1) (Galli 1991, Leblanc et al. 1997).
CHEMICAL EFFECTS OF URBANIZATION
Chemical effects of urbanization are far more variable than hydrologic or geomor-
phic effects and depend on the extent and type of urbanization (residential versus
commercial/industrial), presence of wastewater treatment plant (WWTP) effluent
and/or combined sewer overflows (CSOs), and the extent of stormwater drainage.
Overall, there are more data on water and sediment chemistry in urban streams
than any other aspect of their ecology. This is aided by several very large national
datasets of stream chemistry that focus in whole or in part on urbanization [e.g.,
National Urban Runoff Program (United States), National Water Quality Assess-
mentProgram (USGS 2001),Land-Ocean Interaction Study (UK) (Neal & Robson
2000)].
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In general, there is an increase in almost all constituents, but consistently in
oxygen demand, conductivity, suspended solids, ammonium, hydrocarbons, and
metals, in urban streams (Porcella & Sorenson 1980, Lenat & Crawford 1994,
Latimer & Quinn 1998, USGS 1999). These increases can be attributed to both
WWTP effluent and non–point source (NPS) runoff. Many countries have ac-
complished significant reductions in chemical constituents as a result of adopting
better WWTP technologies (e.g., Krug 1993, Litke 1999). However, treatment
cannot remove all constituents from wastewater, treatment systems fail, and per-
mitted discharge limits are exceeded. There are more than 200,000 discharges
subject to permitting in the United States (USEPA 2001), and of 248 urban cen-
ters studied, 84% discharge into rivers (40% of those into rivers with mean an-
nual discharges less than 28 m3/s) (Heaney & Huber 1984). In addition, CSO
systems are still common, in which stormwater and untreated sewage are com-
bined and diverted to streams and rivers during storms. At least 28% of the
urban centers mentioned above contained CSOs, and in the United Kingdom
35% of the annual pollutant discharge comes from CSOs and storm drains dur-
ing less than 3% of the time (Heaney & Huber 1984, Faulkner et al. 2000). In
addition, illicit discharge connections, leaking sewer systems, and failing sep-
tic systems are a large and persistent contributor of pollutants to urban streams
(Faulkner et al. 2000). In the Rouge River catchment in Detroit, Michigan, the fo-
cus of an intense federal NPS management program, septic failure rates between
17% and 55% were reported from different subcatchments, and it was estimated
thatillicit untreatedsewagedischarge volumeat more than 193,000 m3/yr (Johnson
et al. 1999). The ubiquitous nature of small, NPS problems in urban catchments
has led some to suggest that the cumulative effect of these small problems may be
the dominant source of biological degradation in urban catchments (Duda et al.
1982).
Nutrients and Other Ions
Urbanization generally leads to higher phosphorus concentrations in urban catch-
ments (Omernik 1976, Meybeck 1998, USGS 1999, Winter & Duthie 2000). An
urban effect is most often seen in total phosphorus as a result of increased particle-
associated phosphorus, but dissolved phosphorus levels are also increased (Smart
et al. 1985). In some cases increases in phosphorus can even rival those seen in
agricultural catchments both in terms of concentration and yield (Omernik 1976).
Even an attempt to understand the agricultural contribution to catchment phos-
phorus dynamics in a midwestern catchment discovered that urbanization was a
dominant factor (Osborne & Wiley 1988). Even though urban areas constituted
only 5% of the catchment area and contributed only a small part to the total annual
yield of dissolved phosphorus, urban land use controlled dissolved phosphorus
concentration throughout the year.
Sources of phosphorus in urban catchments include wastewater and fertilizers
(LaValle 1975). Lawns and streets were the primary source of phosphorus to urban
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ECOLOGY OF URBAN STREAMS 343
streams in Madison, Wisconsin as a result of fertilizer application (Waschbusch
et al. 1999). Soils are important in phosphorus dynamics, and the retention of
groundwater phosphorus from septic fields affects stream phosphorus concentra-
tions (Hoare 1984, Gerritse et al. 1995). Phosphorus stored in soils as a result of
fertilization, however, can be mobilized by soil erosion and contribute to eutroph-
ication of receiving waters. This effect has been called the “chemical time bomb”
and is of particular concern when previously agricultural land is cleared for urban
growth (Bennett et al. 1999).
Althoughphosphorus concentrationsare elevatedinurban streams,the effective
increase is not as great as that observed for nitrogen. Urban centers have been
shown to increase the nitrogen concentration in rivers for hundreds of kilometers
(Meybeck1998,USGS 1999).Increaseshavebeen observedforammonium aswell
as nitrate (McConnell 1980, Hoare 1984, Zampella 1994, Wernick et al. 1998).
The extent of the increase depends on wastewater treatment technology, degree
of illicit discharge and leaky sewer lines, and fertilizer use. As with phosphorus,
nitrogen concentrations in streams draining agricultural catchments are usually
much higher (USGS 1999), but some have noticed similar or even greater levels
of nitrogen loading from urbanization (Omernik 1976, Nagumo & Hatano 2000).
Soilcharacteristics also affect the degree of nitrogenretention, of importance when
on-site septic systems are prevalent (Hoare 1984, Gerritse et al. 1995).
Other ions are also generally elevated in urban streams, including calcium,
sodium, potassium, and magnesium (McConnell 1980, Smart et al. 1985,
Zampella 1994, Ometo et al. 2000). Chloride ions are elevated in urban streams,
especially where sodium chloride is still used as the principal road deicing salt.
A significant portion of the more than 100,000 tons of sodium chloride applied in
metropolitan Toronto annually for deicing enters long-turnover groundwater pools
and is released slowly, raising stream chloride concentrations throughout the year
(Howard & Haynes 1993). The combined effect of heightened ion concentrations
instreams isthe elevatedconductivityobservedin most urban streams. Theeffect is
so common that some have suggested using chloride concentration or conductivity
as general urban impact indicators (Wang & Yin 1997, Herlihy et al. 1998).
Metals
Another common feature of urban streams is elevated water column and sedi-
ment metal concentrations (Bryan 1974, Wilber & Hunter 1977, Neal et al. 1997,
Horowitz et al. 1999, Neal & Robson 2000). The most common metals found
include lead, zinc, chromium, copper, manganese, nickel, and cadmium (Wilber
& Hunter 1979), although lead has declined in some urban river systems since its
elimination as a gas additive (Frick et al. 1998). Mercury is also elevated in some
urban streams, and particle-bound methyl-mercury can be high during stormflow
(Mason& Sullivan1998,Horowitzet al.1999). Inaddition toindustrial discharges,
there are many NPSs of these metals in urban catchments: brake linings contain
nickel, chromium, lead, and copper; tires contain zinc, lead, chromium, copper,
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344 PAUL ¥MEYER
and nickel; and metal alloys used for engine parts contain nickel, chromium, cop-
per, and manganese among others (Muschak 1990, Mielke et al. 2000). All of
these metals accumulate on roads and parking lots (Sartor et al. 1974, Forman &
Alexander 1998). Many other metals have been found in elevated concentrations
in urban stream sediments including arsenic, iron, boron, cobalt, silver, strontium,
rubidium,antimony,scandium, molybdenum, lithium, and tin (Khamer etal. 2000,
Neal & Robson 2000). Not surprisingly, it appears that NPSs of metals are more
important than point sources in urban streams (Wilber & Hunter 1977, Mason &
Sullivan 1998).
Theconcentration,storage, and transportofmetals inurbanstreamsis connected
to particulate organic matter content and sediment characteristics (Tada & Suzuki
1982, Rhoads & Cahill 1999). Organic matter has a high binding capacity for
metals, and both bed and suspended sediments with high organic matter content
frequently exhibit 50–7500 times higher concentrations of zinc, lead, chromium,
copper, mercury, and cadmium than sediments with lower organic matter content
(Warren & Zimmerman 1994, Mason & Sullivan 1998, Gonzales et al. 2000).
Sediment texture is also important, and metal concentration in sediments was
inversely correlated to sediment particle size in several urban New Jersey streams
(Wilber & Hunter 1979). In addition, geomorphic features have been shown to
influence metal accumulations. Higher sediment metal concentrations were found
in areas of low velocity (stagnant zones, bars, etc.) where fine sediments and
organic particles accumulate, whereas areas of intermediate velocities promoted
the accumulation of sand-sized metal particles, which can also be common in
urban streams (Rhoads & Cahill 1999).
Several organisms (including algae, mollusks, arthropods, and annelids) have
exhibited elevated metal concentrations in urban streams (Davis & George 1987,
Rauch & Morrison 1999, Gundacker 2000), and ecological responses to metals
include reduced abundances and altered community structure (Rauch & Morrison
1999). It is important to note that the route of entry appears to be both direct expo-
sure to dissolved metals and ingestion of metals associated with fine sediments and
organic matter. This has led a few researchers to suggest that metal toxicity is most
strongly exerted through the riverbed rather than the overlying water (Medeiros
et al. 1983, House et al. 1993), although only dissolved metal concentrations in
the water column are regulated in the United States.
Pesticides
Pesticide detection frequency is high in urban streams and at concentrations fre-
quently exceeding guidelines for the protection of aquatic biota (USGS 1999,
Hoffman et al. 2000). These pesticides include insecticides, herbicides, and fungi-
cides(Daniels et al. 2000). In addition, the frequent detection ofbanned substances
suchas DDT and other organochlorine pesticides (chlordane anddieldrin) in urban
streamsremains a concern (USGS 1999).Most surprising isthat manyorganochlo-
rine pesticide concentrations in urban sediments and biota frequently exceed those
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ECOLOGY OF URBAN STREAMS 345
observed in intensive agricultural areas in the United States (USGS 1999), a phe-
nomenon observed in France as well (Chevreuiel et al. 1999). Additionally, it is
estimated that the mass of insecticides contributed by urban areas is similar to that
from agricultural areas in the United States (Hoffman et al. 2000).
There are many sources of pesticides in urban catchments. Urban use accounts
for more than 136,000 kg, which is a third of U.S. pesticide use (LeVeen & Willey
1983). They are frequently applied around homes (70–97% of U.S. homes use pes-
ticides) and commercial/industrial buildings and are intensively used in lawn and
golf course management (LeVeen & Willey 1983, USGS 1999). Areal application
rates in urban environments frequently exceed those in agricultural applications
by nearly an order of magnitude (Schueler 1994b). For example, pesticide appli-
cation rates on golf courses (including herbicides, insecticides, and fungicides)
exceed 35 pounds/acre/year, whereas corn/soybean rotations receive less than
6 pounds/acre/year (Schueler 1994b). However, unlike agricultural use, urban pes-
ticideapplication rates are generally not well documented(LeVeen&Willey 1983,
Coupe et al. 2000).
As with metals, the main vector of transport of pesticides into urban streams
appears to be through NPS runoff rather than WWTP effluent (Foster et al. 2000).
A strong correlation between particle concentration and pesticide concentration
was found in the Anacostia River basin in Maryland and the San Joaquin River in
California, suggesting NPS inputs are most important (Pereira et al. 1996, Foster
et al. 2000). Volatilization and aerosol formation contributed to higher pesti-
cide concentrations, including atrazine, diazinon, chlorpyrifos, p,p0-DDE (a DDT
metabolite), and other organochlorines, in precipitation in urban areas and may
contribute directly to greater pesticide concentrations and yields in urban areas
(Weibel et al. 1966, Coupe et al. 2000).
Other Organic Contaminants
A whole suite of other organic contaminants are frequently detected in urban
streams, including polychlorinated biphenyls (PCBs), polycyclic aromatic hydro-
carbons (PAHs), and petroleum-based aliphatic hydrocarbons (Whipple & Hunter
1979,Moring & Rose 1997,Frick et al. 1998).PCBs are still frequently detected in
urban areas of the United States, even though their use in manufacturing was out-
lawed because of their carcinogenic effects. These compounds are very stable and
are still found in fish at concentrations exceeding consumption-level guidelines
in urban rivers such as the Chattahoochee River below Atlanta, Georgia (Frick
et al. 1998). PCB concentrations were highly correlated with urban land use in
the Willamette Basin in Oregon as well (Black et al. 2000). As with metals and
pesticides, PCBs are primarily particle associated, and in the absence of industrial
point sources, it is assumed that stormwater runoff is the major route of entry
(Foster et al. 2000).
PAHs are a large class of organic compounds that include natural aromatic
hydrocarbons but also many synthetic hydrocarbons including organic solvents
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346 PAUL ¥MEYER
with different industrial uses (Yamamoto et al. 1997). For this reason, the unnat-
ural PAHs are probably derived from industrial effluent or episodic spills. Very
little is known about these compounds in urban streams. In Dallas–Fort Worth,
Texas streams, 24 different industrial PAHs were detected, including 4 of the
top 10 U. S. Environmental Protection Agency (EPA) most hazardous substances,
and at concentrations exceeding human health criteria (Moring & Rose 1997).
In Osaka, Japan streams, 55 PAHs were detected, including 40 EPA target com-
pounds. Organic solvents (e.g., toluene, trichloroethane, and dichlorethane) were
most common (Yamamoto et al. 1997).
It is difficult to find automobile parking spaces without oil stains in any city.
The result of these leaky crankcases is a cornucopia of different petroleum-
based aliphatic hydrocarbons in storm runoff associated primarily with particles
(Whipple & Hunter 1979). Although there are natural aliphatic hydrocarbons in
streams, these are generally overwhelmed by petroleum-based compounds in ur-
ban stream bed and water-column sediments (Hunter et al. 1979, Mackenzie &
Hunter 1979, Eganhouse et al. 1981). Evidence suggests that these are frequently
at concentrations that are stressful to sensitive stream organisms (Latimer & Quinn
1998). Most striking is the yield of these compounds from urban catchments. An
estimated 485,000 liters of oil enters the Narragansett Bay each year, a volume
equal to nearly 50% of the disastrous 1989 World Prodigy oil spill in that same
bay (Hoffman et al. 1982, Latimer & Quinn 1998). Similarly, it is estimated that
the Los Angeles River alone contributes about 1% of the annual world petroleum
hydrocarbon input to the ocean (Eganhouse et al. 1981).
Lastly, recent data suggest pharmaceutical substances from hospital effluent
may contribute an array of different chemical compounds into streams. Detectable
levelsofantiobiotics,genotoxic chemotherapeuticdrugs,analgesics, narcotics,and
psychotherapeutic drugs have been reported from effluent and/or surface waters
(Halling-Sorensen et al. 1998). Although there is some information on the toxicity
of these different compounds from laboratory studies, there are insufficient data
on the nature or extent of the threat they pose to urban stream biota.
BIOLOGICAL AND ECOLOGICAL EFFECTS
OF URBANIZATION
The ecological implications of urbanization are far less studied than the chemical
effects, an absence noted in several studies (Porcella & Sorenson 1980, Duda et al.
1982, Medeiros et al. 1983). Nevertheless, much is known about the response of
streamorganisms, especially invertebrates,to urbanization;far less is known about
urban effects on fish (Mulholland & Lenat 1992). Of even greater concern is the
lack of mechanistic studies; few studies analyze whether physical habitat, water
quality, or food web disturbances (either resource effects or altered community
interactions) are the cause of biological degradation in urban streams (Suren
2000). Grossly underrepresented are studies of population dynamics, community
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ECOLOGY OF URBAN STREAMS 347
interactions, and ecosystem ecology of urban streams, which is surprising given
the level of knowledge within the field (Allan 1995). Lastly, very little information
has been gathered on biological monitoring of restoration or best management
practice implementation in urban catchments (Riley 1998). Most studies assess
performance based on stream channel condition or pollutant reduction; few, if any,
monitor biological response (Benke et al. 1981, Center for Watershed Protection
2000). In this section, we discuss the effects of urbanization on microbes, algae,
macrophytes, invertebrates, and fish.
Microbes
Bacterial densities are usually higher in urban streams, especially after storms
(Porcella & Sorenson 1980, Duda et al. 1982). Much of this is attributable to
increased coliform bacteria, especially in catchments with wastewater treatment
plant (WWTP) and combined sewer overflow (CSO) effluent (Gibson et al. 1998,
Young & Thackston 1999). In Saw Mill Run, an urban stream near Pittsburgh,
Pennsylvania, fecal coliform colony–forming units (CFU) increased from 170–
13,300 CFU/100 ml during dry weather to 6,100–127,000 CFU/100 ml during wet
weather (Gibson et al. 1998). CSOs contributed 3,000–85,000 CFU/100 ml during
wet weather. These data indicate that non–point sources (NPSs) as well as point
sources contribute to fecal coliform loads in urban streams. High values during dry
weather are not uncommon in urban streams and may indicate chronic sewer leak-
age or illicit discharges. Storm sewers were also a significant source of coliform
bacteria in Vancouver, British Columbia; stormwater there contained both human
and nonhuman fecal coliform bacteria (Nix et al. 1994). Other pathogens, includ-
ing Cryptosporidum and Giardia, have also been associated with CSOs (Gibson
et al. 1998).
Increased antibiotic resistance has been seen in some urban bacterial popula-
tions (Goni-Urriza et al. 2000). Increased resistance to several antiobiotics, in-
cluding nalidixic acid, tetracycline, beta-lactam, and co-trimoxazole, has been
observed from several enteric as well as native stream species isolated from a
river downstream of a WWTP discharge in Spain. It may be that resistant bacte-
ria are passing through the treatment process and conferring resistance to native
bacteria. Recent evidence suggests that metal toxicity may also be indirectly in-
volved in increasing antibiotic resistance in stream bacteria. Bacterial resistance
to streptomycin and kanamycin were positively correlated with sediment mercury
concentration in streams below nuclear reactors and industrial facilities, a result
of indirect selection for metal tolerance (McArthur & Tuckfield 2000). Metals
may also affect bacterial enzyme activity in urban streams. Enzyme levels were
inversely correlated to sediment metal concentration in an urban stream, and this
was especially pronounced below an industrial effluent (Wei & Morrison 1992).
Nitrifying bacteria, responsible for the oxidation of reduced nitrogen, are also
influenced by urbanization. WWTP effluent can represent a significant source of
nitrifying bacteria to urban streams (Brion & Billen 2000). These bacteria are
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348 PAUL ¥MEYER
used to oxidize ammonium during the treatment process, but escape into streams
ineffluent and contributeto the high nitrifieractivity observed below some WWTP
discharges (Jancarkova et al. 1997). Nitrification rates were as much as six times
higher in treated effluent entering the Seine than in receiving river water upstream
(Brion&Billen 2000). Ironically,becausesomanynitrifiersentered theSeineRiver
in France via untreated sewage historically, the reduction in untreated sewage via
improvedsewagedesign contributed toa reductionin ammoniumoxidation rates in
the river from 1.5 µmol/liter/h in 1976 to 1.0 µmol/liter/h in 1993 (Brion & Billen
2000). In addition to nitrifiers, iron-oxidizing bacteria are often abundant in urban
streams, especially where reduced metals emerge from anoxic urban groundwater
or storm sewers (Dickman & Rygiel 1998).
Algae
Theuse ofalgae to indicate water qualityin Europeand the United States hasa long
history(Kolkwitz&Marsson 1908, Patrick1973).As a result,information exists on
algalspecies and communityresponses to organicpollution; however, theresponse
of algae to all aspects of urbanization is far less studied. The increasing proportion
ofurban land usein acatchment generallydecreases algal speciesdiversity,andthis
change has been attributed to many factors including water chemistry (Chessman
etal. 1999). Elevatednutrientsand light levelstypicallyfavorgreateralgal biomass,
which has been observed in many urban streams, where algae do not appear to
be nutrient limited (Chessman et al. 1992, Richards & Host 1994). However, the
shifting nature of bed sediment in urban streams, frequent bed disturbance, and
high turbidity may limit algal accumulation (Burkholder 1996, Dodds & Welch
2000).In addition, severalalgalspecies aresensitivetometals, andstream sediment
metalaccumulation canresult inreduced algal biomass (Olguin et al. 2000).Lastly,
the frequent detection of herbicides in streams, some with known effects on algae
(Davies et al. 1994), will undoubtedly affect stream algal communities
Macrophytes
Little has been written on macrophyte response to urbanization. Most of the work
hasbeen donein NewZealand and Australia, wherebed sedimentchanges, nutrient
enrichment,and turbidity allcontribute toreduced diversityof stream macrophytes
(Suren 2000). Exotic species introductions in urban streams have also resulted
in highly reduced native macrophyte diversity (Arthington 1985, Suren 2000).
Excessive macrophyte growth as a result of urbanization has not been observed in
New Zealand, even though nutrient and light levels are higher (Suren 2000).
Invertebrates
Literature searches revealed more studies of urban effects on aquatic invertebrates
than on any other group, and the available data are being expanded by groups
biomonitoring urban systems (e.g., USGS National Water Quality Assessment,
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ECOLOGY OF URBAN STREAMS 349
U.S. EPA, state agencies, and others). All aspects of aquatic invertebrate habitat
are altered by urbanization. One of the historically well-studied aspects has been
the effects of organic pollutants (especially WWTP effluent) on invertebrates. Or-
ganic pollution generally reduces invertebrate diversity dramatically, resulting in
a community dominated by Chironomidae (Diptera) and oligochaetes (Campbell
1978, Seager & Abrahams 1990, Wright et al. 1995). However, general effects
of urbanization on stream invertebrates have also been studied and general in-
vertebrate responses can be summarized as follows: decreased diversity in re-
sponse to toxins, temperature change, siltation, and organic nutrients; decreased
abundances in response to toxins and siltation; and increased abundances in re-
sponse to inorganic and organic nutrients (Resh & Grodhaus 1983, Wiederholm
1984).
Studiesof the effectsof urbanland use oninvertebrates can be dividedinto three
types: those looking along a gradient of increasing urbanization in one catchment,
those looking at an urbanized versus a reference catchment, and large studies
considering urban gradients and invertebrate response in several catchments. All
single catchment gradient studies find a decrease in invertebrate diversity as ur-
ban land use increases, regardless of the size of the catchment (Pratt et al. 1981,
Whiting & Clifford 1983, Shutes 1984, Hachmoller et al. 1991, Thorne et al.
2000). Decreases were especially evident in the sensitive orders—Ephemeroptera,
Plecoptera, and Trichoptera (Pratt et al. 1981, Hachmoller et al. 1991). Most of
these studies observed decreases in overall invertebrate abundance, whereas the
relative abundance of Chironomidae, oligochaetes, and even tolerant gastropods
increased (Pratt et al. 1981, Thorne et al. 2000). Comparative catchment studies
show the same trends with increasing urbanization as those observed in single
catchment studies: decreased diversity and overall abundance and increased rela-
tive abundance of tolerant Chironomidae and oligochaetes (Medeiros et al. 1983,
Garie & McIntosh 1986, Pederson & Perkins 1986, Lenat & Crawford 1994).
Themulti-catchment studies attempt to relate differing amountsof urbanization
in many catchments to particular invertebrate community responses, often using a
gradient analysis approach. As discussed above, all find decreases in diversity and
overall invertebrate abundance with increased urbanization. This response is cor-
related with impervious surface cover, housing density, human population density,
and total effluent discharge (Klein 1979, Benke et al. 1981, Jones & Clark 1987,
Tate & Heiny 1995, Kennen 1999). Klein (1979) studied 27 small catchments on
the Maryland Piedmont and was among the first to identify impervious surface
cover (ISC) as an important indicator of degradation. Invertebrate measures de-
clinedsignificantly with increasing ISCuntil they indicatedmaximum degradation
at 17% ISC (Table 1). Degradation thresholds at ISC between 10 and 20% have
been supported by numerous other studies for many different response variables
(see Schueler 1994a). Residential urbanization in Atlanta, Georgia had dramatic
effects on invertebrate diversity, but there were very few clues as to the mecha-
nisms responsible, although leaky sewers were implicated in these and other urban
residential catchments (Benke et al. 1981, Johnson et al. 1999).
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350 PAUL ¥MEYER
Few studies have considered specific mechanisms leading to the observed ef-
fects of urbanization. This is a difficult task because of the multivariate nature of
urban disturbance. Increased turbidity has been associated with higher drift densi-
tiesofinsects (Doeg &Milledge1991), but more workhasfocused ontheinstability
of smaller and more mobile bed sediments associated with urban sedimentation. In
general, the change in bed sediments favors species adapted to unstable habitats,
such as the chironomid dipterans and oligochaete annelids (Pedersen & Perkins
1986, Collier 1995). Where slopes are steeper, and smaller sediments are removed
by increased water velocities, localized areas of higher invertebrate diversity are
observed within the coarser sediments (Collier 1995). Pools are particularly af-
fected by sediment accumulation in urban streams, and invertebrate communities
within these habitats are degraded (Hogg & Norriss 1991). Lastly, sedimentation
associated with urban streams reduces available refugial space, and invertebrates
are more susceptible to drift when refugial space is limited during the frequent
floods characteristic of urban environments (Borchardt & Statzner 1990). Storm-
flows in urban streams introduce the majority of pollutants and also move the bed
sediment frequently. The mortality of Pteronarcys dorsata (Plecoptera) in cages in
urban streams was attributed to sedimentation associated with storms (Pesacreta
1997).
Sediment toxicity has also been explored. As mentioned above, benthic organic
matter binds many toxins and is also a major food resource for many stream
invertebrates (Benke & Wallace 1997). Mortality of aquatic invertebrates remains
high in many urban streams even during low flow periods, suggesting that toxicity
associated with either exposure in the bed or ingestion of toxins associated with
organic matter contributes to invertebrate loss (Pratt et al. 1981, Medeiros et al.
1983).
Riparian deforestation associated with urbanization reduces food availability,
affects stream temperature, and disrupts sediment, nutrient, and toxin uptake from
surface runoff. Invertebrate bioassessment metrics decreased sharply in Puget
Sound, Washington tributaries with increasing ISC (Horner et al. 1997). However,
streams that had higher benthic index of biotic integrity scores for a given level of
ISC were always associated with greater riparian forest cover in their catchment,
suggesting that riparian zones in some urban catchments may buffer streams from
urban impacts. Above 45% ISC, all streams were degraded, regardless of riparian
status. The value of riparian forests is also reduced if the stormwater system is
designed to bypass them and discharge directly into the stream.
Road construction associated with urbanization impacts stream invertebrates.
Long-term reductions (>6 y) in invertebrate diversity and abundances were ob-
served in association with a road construction project in Ontario (Taylor & Roff
1986). General effects of roads on streams has been reviewed recently (Forman &
Alexander 1998).
Very little ecological data beyond presence/absence or abundance data have
been reported for urban stream invertebrates. Aquatic insect colonization potential
wasreported tobe highin someurban streams,suggesting restorationefforts would
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ECOLOGY OF URBAN STREAMS 351
not be limited in this regard (Pedersen & Perkins 1986), but little is known about
colonizationor adult aquaticinsect ecologyin urban streams.Urban streamrestora-
tion work focuses largely on channel geomorphological stability, with relatively
little attention given to biological restoration (Riley 1998), although restoration
of Strawberry Creek on the campus of the University of California at Berkeley
has resulted in detectable increases in invertebrate diversity and abundance
(Charbonneau & Resh 1992). Drift of aquatic invertebrates is a well studied phe-
nomenon in streams, but with one exception (Borchardt & Statzner 1990), little
has been published on insect drift in urban streams. We found no published work
regarding life cycle ecology (e.g., voltinism or emergence timing), population
dynamics, behavioral ecology, community interactions, or production of aquatic
invertebrates in urban streams.
Fish
Less is known about fish responses to urbanization than about invertebrates, and
a general response model does not exist. However, the Ohio Environmental Pro-
tection Agency has a very large database of land use and fish abundance from
around their state and has suggested three levels of general fish response to in-
creasing urbanization: from 0 to 5% urban land use, sensitive species are lost; from
5 to 15%, habitat degradation occurs and functional feeding groups (e.g., benthic
invertivores) are lost; and above 15% urban land use, toxicity and organic enrich-
ment result in severe degradation of the fish fauna (Table 1) (Yoder et al. 1999).
This model has not been verified for other regions of the country, where studies
have focused on various aspects of urbanization. Here we consider three types of
urban land use studies with regards to fish: gradients of increasing urbanization
withina single catchment, comparingan urban and reference catchment, and large,
multi-catchment urban gradient studies.
Along urban gradients within single catchments, fish diversity and abundances
decline, and the relative abundance of tolerant taxa increases with increasing ur-
banization (Table 1) (Onorato et al. 2000, Boet et al. 1999, Gafny et al. 2000).
Invasive species were also observed to increase in more urbanized reaches of the
Seine River, France, and this effect extended more than 100 km below Paris (Boet
et al. 1999). Summer storms in that river were associated with large fish kills as
a result of dissolved oxygen deficits, an effect also observed for winter floods in
Yargon Stream, the largest urban stream in Israel (Gafny et al. 2000). Comparisons
with historical collections, an approach used commonly with fish studies, revealed
that several sensitive species were extirpated from the Upper Cahaba River system
in Alabama between 1954 and 1995, a period coinciding with the rapid growth
of Birmingham, Alabama (Onorato et al. 2000). Extirpation of fish species is not
uncommon in urban river systems (Ragan & Dietmann 1976, Weaver & Garman
1994, Wolter et al. 2000).
Comparative catchment studies also find dramatic declines in fish diversity and
abundances in urban catchments compared with forested references (Scott et al.
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352 PAUL ¥MEYER
1986, Weaver & Garman 1994, Lenat & Crawford 1994). Kelsey Creek, a well-
studied urban stream in Washington, is unusual in that it has sustained salmonid
populations, especially cutthroat trout (Oncorhynchus clarki), even though coho
salmon (Oncorhyncus kisutch) and many nonsalmonid species have disappeared
(Scott et al. 1986). Salmonids in the urban stream actually grow more rapidly
and to larger sizes, increasing fish production up to three times that in the forested
referencesite, presumably aresult of warmertemperatures and greaterinvertebrate
biomass in the urban stream. However, the population size structure is different in
the two streams, with year 0 and 1 cutthroat underrepresented in the urban stream
(Scott et al. 1986).
Large multi-site studies of fish responses to urban gradients also find dramatic
decreases in diversity or fish multimetric indices [index of biotic integrity (IBI)]
with increasing ISC or other urban land use indicators (Table 1) (Klein 1979,
Steedman 1988, Wang et al. 1997, Frick et al. 1998, Yoder et al. 1999). Similar to
effects observed for invertebrates, these studies also find precipitous declines in
fish metrics between 0 and 15% ISC or urban land use, beyond which fish commu-
nities remain degraded (Klein 1979, Yoder et al. 1999). The effect of urbanization
on fish appears at lower percent land area disturbed than effects associated with
agriculture. In Wisconsin and Michigan few fish community effects were observed
in agricultural catchments up to 50% agricultural land use in the catchment (Roth
et al. 1996, Wang et al. 1997), and mixed agriculture and urban catchments had
significantly lower IBI scores than strictly agricultural catchments (Wang et al.
2000). This suggests that although total urban land use occupies a smaller area
globally, it is having disproportionately large effects on biota when compared with
agriculture. However, it is crucial to recognize that all urban growth does not have
the same effects. Extensive fish surveys in Ohio suggest that residential develop-
ment, especially large-lot residential development, has less of an effect on stream
fishes than high-density residential or commercial/industrial development (Yoder
et al. 1999). They hypothesize that riparian protection and less channel habitat
degradation are responsible for protecting the fauna in these streams, even up to
15% urban land use. Similar benefits of riparian forests to fish in urban streams
were observed in the Pacific Northwest (Horner et al. 1997).
Few studies have explored specific mechanisms causing changes in fish assem-
blages with urbanization. Sediment is presumably having effects on fish in urban
streams similar to those observed in other systems although toxin-mediated im-
pacts may be greater (Wood & Armitage 1997). Road construction results in an
increase in the relative abundance of water-column feeders as opposed to benthic
feeders, likely a response to a decrease in benthic invertebrate densities (Taylor
& Roff 1986). Benthic feeders quickly reappeared as sedimentation rates de-
clined after construction. Flow modification associated with urbanization also
affects stream fish. In the Seine, modification of flow for flood protection and
water availability has affected pike (Esox lucius) by reducing the number of
flows providing suitable spawning habitat. With urbanization, the river contains
enough suitable spawning habitat in only 1 out of 5 years as opposed to 1 out of
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ECOLOGY OF URBAN STREAMS 353
every 2 years historically (Boet et al. 1999). Last, WWTP effluent clearly affects
fishes.Reductions in WWTP effluenthavebeen associated with therecovery ofthe
fish community in a River Trent tributary near Birmingham, England (Harkness
1982). After nearly 250 years of degradation, effluent reductions, improved treat-
ment, and construction of run-of-the river purification have resulted in an increase
in fish diversity and abundances.
A few studies have actually examined ecological factors regulating stream fish
populations and communities in urban streams. Recruitment of anadromous fish
in the Hudson River Basin in New York was limited by suitable spawning habitat
as a result of urbanization (Limburg & Schmidt 1990). Numbers of alewife (Alosa
pseudoharengus) eggs and larvae in tributary streams decreased sharply between
0 and 15% urban land use. Beyond 15%, no eggs or larvae were found. The Kelsey
Creek study discussed above showed impacts on salmonid population structure
associated with urbanization, suggesting that urban streams may serve as popula-
tion sinks for cutthroat, and that fish populations in those streams are dependent on
recruitment from source populations with normal population age structures (Scott
et al. 1986). Few data on the diet of fish in urban streams have been published,
although a shift in diet was observed for fish along an urban gradient in Virginia
(Weaver & Garman 1994).
Introduced fish species are also a common feature of urban streams. As a result
of channelization, other river transportation modifications, and voluntary fisheries
efforts in the Seine around Paris, 19 exotic species have been introduced, while
7 of 27 native species have been extirpated (Boet et al. 1999). The red shiner
(Cyprinella lutrensis), a Mississippi drainage species commonly used as a bait
fish, has invaded urban tributaries of the Chattahoochee River in Atlanta, Georgia
where it has displaced native species and now comprises up to 90% of the fish
community (DeVivo 1995).
As observed above for invertebrates, real gaps exist in our understanding of
fish ecology in urban streams. The effects of urbanization on fishes have focused
primarilyon patterns ofspecies presence,absence, orrelativeabundance.Wefound
no published information on behavioral ecology, community interactions, or the
biomass and production of nonsalmonid fishes in urban streams.
Ecosystem Processes
Ecosystem processes such as primary productivity, leaf decomposition, or nutrient
cycling have been overlooked in urban streams, although they have been exten-
sivelystudied in other typesof stream ecosystems(Allan 1995). A fewstudies have
consideredorganic matter in streams.WWTP effluentand CSO dischargescan dra-
matically increase dissolved and particulate organic carbon concentrations, espe-
cially during storms (McConnell 1980). However, much less is known about base-
flow concentrations of particulate and dissolved carbon in urban streams—natural
or anthropogenic. The carbon inputs associated with sewage are generally more
labile than natural transported organic matter and they affect dissolved oxygen
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354 PAUL ¥MEYER
in streams. Oxygen deficits associated with high biological oxygen demand dur-
ing and after storms are common (McConnell 1980, Faulkner et al. 2000, Ometo
et al. 2000). In addition, nonrespiratory oxygen demands associated with chemi-
cal oxidation reactions are also elevated in urban streams and can be much higher
than biological oxygen demand in stormwater runoff (Bryan 1972). These inputs
explain in part why more than 40% of 104 urban streams studied in the United
States showed a high probability of greater than average oxygen deficits, with
dissolved oxygen concentrations below 2 mg/liter and daily fluctuations up to
7 mg/liter not uncommon (Keefer et al. 1979). In a comparison of 2 forested and
4 urban catchments, average organic matter standing stocks were significantly
lower in urban streams near Atlanta, Georgia (Paul 1999). This was attributed to
greater scouring of the highly mobile sandy substrates in urban channels as a result
of more severe flows.
Organic matter quality has been characterized in a few urban streams. In Kelsey
Creek, particulate organic matter (POM) carbohydrate concentrations were higher
than in POM in a nearby forested reference stream, suggesting that urbanization
affectsthe nature oftransported organic matter as well (Sloane-Richeyet al. 1981).
Inaddition to differencesin organic matter quantity and quality,urban streams also
differ in organic matter retention. Coarse and fine particles released to measure
organic matter transport in Atlanta, Georgia streams traveled much farther before
leaving the water column in urban streams than in forested streams (Paul 1999).
Combined with the data from benthic organic matter (BOM) storage, these data
indicate that these urban streams retain less organic matter, a fact that could limit
secondary production in these urban streams (Paul 1999).
Ecosystem metabolism has also been measured in a few urban streams. In a
comparison of three rivers in Michigan the urban river had higher gross primary
production and community respiration than the forested river (Ball et al. 1973). In
addition, the gross primary productivity to community respiration (P/R) ratio in
the urban river without municipal effluent was greater than the forested stream and
greater than 1.0, indicating that autotrophy dominated organic matter metabolism.
However, in a downstream reach of the urban river receiving effluent, respiration
was higher and the P/R ratio less than the forested river and far less than 1.0, indi-
cating that heterotrophic metabolism predominated. Similar results were observed
for urban streams in Atlanta, where gross primary production and community res-
piration were higher in urban streams than forested streams, and urban streams had
more negative net ecosystem metabolism (gross primary production–community
respiration),indicating greaterheterotrophy (Paul1999). However,because carbon
storage was far less in the urban streams, carbon turnover was faster, supporting
the hypothesis that respiration in urban streams was driven by more labile sources
of carbon, such as sewage effluent.
Decomposition of organic matter has been measured in a few urban streams.
Willow leaves decayed much faster in two suburban New Zealand streams than
ever reported for any other stream; this occurred regardless of whether shredding
insectswerepresent orabsent(Collier &Winterbourn1986).The sameresultswere
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ECOLOGY OF URBAN STREAMS 355
observedforchalk maple (Acerbarbatum)decay inurbanstreams inAtlanta,where
rateswere farfaster in urban streamsthan ratesobserved for any woody leafspecies
in any stream (Paul 1999). Fungal colonization of leaves was only slightly lower in
theurban streams,but there were noshredding insectsassociated withpacks. These
results suggested that higher stormflow was responsible for greater fragmentation
of leaves in the urban streams, resulting in faster decay rates (Paul 1999).
Removal of added nutrients and contaminants is an ecological service provided
by streams and relied upon by society. Although nutrient uptake in flowing waters
has been extensively studied in forested ecosystems (Meyer et al. 1988, Stream
Solute Workshop 1990, Marti & Sabater 1996), urban settings have been largely
ignored. Studies in enriched reaches of river below the effluent from wastewater
treatment plants have provided opportunities to examine patterns of denitrification
in rivers (e.g., Hill 1979) and seasonal patterns of phosphorus removal and reten-
tion in a eutrophic river (e.g., Meals et al. 1999). Recently, ecologists have used the
nutrients added by a wastewater treatment plant to measure nutrient uptake length,
which is the average distance downstream traveled by a nutrient molecule before it
is removed from the water column (Marti et al. 2001, Pollock & Meyer 2001). Up-
take lengths in these rivers are much longer than in nonurban rivers of similar size,
suggestingthat notonly isnutrient loadingelevatedin urban streams, but alsonutri-
entremovalefficiencyis greatlyreduced. Thenet resultof thesealterations inurban
streamsisincreased nutrientloading todownstreamlakes,reservoirs, and estuaries.
OPPORTUNITIES AND IMPERATIVES FOR
AN ECOLOGY OF URBAN STREAMS
Urban streams are common features of the modern landscape that have received
inadequateecological attention.That isunfortunate becausethey offer afertile test-
ing ground for ecological concepts. For example, hydrologic regime is a master
variable in streams (Minshall 1988), influencing channel form, biological assem-
blages, and ecosystem processes. As discussed in this review, impervious surfaces
resultincharacteristically alteredand often extremehydrologic conditions thatpro-
vide an endpoint on a disturbance gradient and that offer opportunities to quantify
the relationships between channel form, biological communities, and ecosystem
processes (Meyer et al. 1988). Does a continuous gradient of impervious surface
cover result in a similar gradient of ecological pattern and process or are there
thresholds? Answering that question is of both theoretical and practical interest.
Developing a mechanistic understanding of the linkages between urbanization and
stream ecosystem degradation is elusive but essential if ecologists hope to under-
standthe nature ofecological response to disturbance and if they wantto contribute
to the development of scenarios that can guide planning decisions.
Many urban centers developed around rivers, which were the lifeblood of com-
merce. These commercial uses of rivers ignored and degraded the ecological ser-
vices rivers provide, a phenomenon continuing today as urban sprawl accelerates.
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356 PAUL ¥MEYER
Despite widespread degradation, urban rivers and streams offer local communities
an easily accessible piece of nature. Most people live in urban areas, and many
children first encounter nature playing in urban streams. Hence, urban streams
offer opportunities for ecological outreach and education that ecologists are only
beginning to explore. The meteoric rise in numbers of local catchment associa-
tions and adopt-a-stream monitoring groups is testimony to an audience eager for
ecological insights.
Urban streams also offer ecologists an opportunity to test concepts of system
organization through restoration projects. The field of urban stream restoration is
dominatedbyphysical scientists andengineersand rarely extendsbeyondstormwa-
ter management and bank stabilization with a goal of reestablishing a channel ge-
omorphology in dynamic equilibrium with the landscape (e.g., Riley 1998). Little
attention is given to restoration of a native stream biota or the ecological services
streams provide. Urban stream restoration offers challenges not only in integrating
physical, chemical, and biological processes to rehabilitate impaired ecosystems,
butalso requires an attention to estheticsand humanattitudes towardthelandscape.
Thisoffers an opportunity for the integration ofecological and social sciences with
landscape design, which if successful will provide an avenue for ecologists to par-
ticipate in the creation of the sustainable metropolitan centers of the future.
Cities have been a part of human history for millenia, and projections suggest
most humans will live in cities in the future. Hence, urban areas lie at the intersec-
tion of human and ecological systems. If we are to succeed in that often-stated goal
of incorporating humans as components of ecosystems, cities and their streams
can no longer be ignored.
ACKNOWLEDGMENTS
This work is dedicated to those who have braved the urban stream. We apologize
to those many the restrictions in length prohibited us from including. Our research
on urban streams in Atlanta has been supported by the EPA/NSF Waters and
Watersheds Program (EPA R 824777-01-0) for work on the Chattahoochee River
and by the EPA Ecological Indicators program (EPA R 826597-01-0) for work on
the Etowah River.
Visit the Annual Reviews home page at www.AnnualReviews.org
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October 9, 2001 11:23 Annual Reviews AR142-FM
Annual Review of Ecology and Systematics
Volume 32, 2001
CONTENTS
CHEMICAL DETECTION OF NATURAL ENEMIES BY ARTHROPODS:
ANECOLOGICAL PERSPECTIVE,Marcel Dicke and Paul Grostal 1
SEX CHROMOSOME MEIOTIC DRIVE,John Jaenike 25
ENVIRONMENTAL INFLUENCES ON REGIONAL DEEP-SEA SPECIES
DIVERSITY,Lisa A. Levin, Ron J. Etter, Michael A. Rex, Andrew J.
Gooday, Craig R. Smith, Jes´
us Pineda, Carol T. Stuart, Robert R. Hessler,
and David Pawson 51
THE PHYSIOLOGY OF LIFE HISTORY TRADE-OFFS IN ANIMALS,
Anthony J. Zera and Lawrence G. Harshman 95
URBAN ECOLOGICAL SYSTEMS:LINKING TERRESTRIAL ECOLOGICAL,
PHYSICAL,AND SOCIOECONOMIC COMPONENTS OF METROPOLITAN
AREAS,S. T. A. Pickett, M. L. Cadenasso, J. M. Grove, C. H. Nilon,
R. V. Pouyat, W. C. Zipperer, and R. Costanza 127
DISPERSAL IN FRESHWATER INVERTEBRATES,David T. Bilton,
Joanna R. Freeland, and Beth Okamura 159
APPLIED EVOLUTION,J. J. Bull and H. A. Wichman 183
MISTLETOE–A KEYSTONE RESOURCE IN FORESTS AND WOODLANDS
WORLDWIDE,David M. Watson 219
THE ROLE OF DISTURBANCE IN THE ECOLOGY AND
CONSERVATION OF BIRDS,Jeffrey D. Brawn, Scott K. Robinson,
and Frank R. Thompson III 251
APPROACHES TO THE STUDY OF TERRITORY SIZE AND SHAPE,
Eldridge S. Adams 277
THE POPULATION BIOLOGY OF INVASIVE SPECIES,Ann K. Sakai,
Fred W. Allendorf, Jodie S. Holt, David M. Lodge, Jane Molofsky,
Kimberly A. With, Syndallas Baughman, Robert J. Cabin, Joel E. Cohen,
Norman C. Ellstrand, David E. McCauley, Pamela O’Neil,
Ingrid M. Parker, John N. Thompson, and Stephen G. Weller 305
STREAMS IN THE URBAN LANDSCAPE,Michael J. Paul and Judy L. Meyer 333
INTEGRATING FUNCTION AND ECOLOGY IN STUDIES OF ADAPTATION:
INVESTIGATIONS OF LOCOMOTOR CAPACITY AS A MODEL SYSTEM,
Duncan J. Irschick and Theodore Garland Jr. 367
v
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October 9, 2001 11:23 Annual Reviews AR142-FM
vi CONTENTS
THE SOUTHERN CONIFER FAMILY ARAUCARIACEAE:HISTORY,
STATUS,AND VALUE FOR PALAEOENVIRONMENTAL
RECONSTRUCTION,Peter Kershaw and Barbara Wagstaff 397
THE UNITS OF SELECTION ON MITOCHONDRIAL DNA, David M. Rand 415
EVOLUTIONARY PATTERNS AMONG PERMO-TRIASSIC THERAPSIDS,
Bruce S. Rubidge and Christian A. Sidor 449
ECOLOGY,CONSERVATION,AND PUBLIC POLICY,Donald Ludwig,
Marc Mangel, and Brent M. Haddad 481
MALE-KILLING,NEMATODE INFECTIONS,BACTERIOPHASE INFECTION,
AND VIRULENCE OF CYTOPLASMIC BACTERIA IN THE GENUS
WOLBACHIA,Lori Stevens, Rosanna Giordano, and Roberto F. Fialho 519
BIOSPHERIC TRACE GAS FLUXES AND THEIR CONTROL OVER
TROPOSPHERIC CHEMISTRY,Russell K. Monson and Elisabeth A.
Holland 547
INDEXES
Subject Index 577
Cumulative Index of Contributing Authors, Volumes 28–32 605
Cumulative Index of Chapter Titles, Volumes 28–32 608
ERRATA
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... Water bodies in densely populated lowland areas are often impacted by multiple stressors (Allan et al., 1997;Paul and Meyer, 2001;Rico et al., Science of the Total Environment 844 (2022) 157045 2016). Hydromorphological alterations cause flow changes and habitat degradation, while excess nutrients, and contaminants like pharmaceuticals, metals, polycyclic aromatic hydrocarbons (PAHs), pesticides, and mixtures of other (un)known compounds originating from human activities, pose chemical stress on these lowland water bodies (Bernhardt et al., 2017;Brown et al., 2009;Paul and Meyer, 2001;Waite et al., 2019). ...
... Water bodies in densely populated lowland areas are often impacted by multiple stressors (Allan et al., 1997;Paul and Meyer, 2001;Rico et al., Science of the Total Environment 844 (2022) 157045 2016). Hydromorphological alterations cause flow changes and habitat degradation, while excess nutrients, and contaminants like pharmaceuticals, metals, polycyclic aromatic hydrocarbons (PAHs), pesticides, and mixtures of other (un)known compounds originating from human activities, pose chemical stress on these lowland water bodies (Bernhardt et al., 2017;Brown et al., 2009;Paul and Meyer, 2001;Waite et al., 2019). Multiple stressors can negatively affect both structure and functioning of aquatic ecosystems (de Vries et al., 2019;, but understanding which (combinations of) stressors actually cause the observed deterioration of ecological integrity remains challenging (Clements, 2000;dos Reis Oliveira et al., 2020;Martinez-Haro et al., 2015). ...
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... Dredging of the riverbed can cause sediment transport from urban and agricultural activities, narrowing of the channel (erosion), degradation of the bed, alteration of the river flow, and reduction or increase in sandbanks and islands, leading to a decrease in diversity of fish species (Taylor et al., 2008). According to Paul and Meyer (2008), the most consistent and widespread effect for sediment transport is the increase in impermeable surface coverage within urban basins, which alters the hydrology and geomorphology of water bodies. ...
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... This translates into more frequent stormflow events with high peak discharge and rapid stormflow recession (flashiness). Urbanisation brings about the redistribution of water from periods of baseflow to periods of stormflow, as well as increased daily variation in streamflow [4,38,[52][53][54][55][56][57][58][59]. Impervious surfaces in immediate riparian zones also increase the risk of stream impairment (due to the decrease in buffer capacity for filtering impaired surface and groundwater) [60]. ...
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