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Streams in the Urban Landscape



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
Michael J. Paul1and Judy L. Meyer
Institute of Ecology, University of Georgia, Athens, Georgia 30602;
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.
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
0066-4162/01/1215-0333$14.00 333
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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
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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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
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.
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
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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.
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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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
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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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).
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 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
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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.
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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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).
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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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.
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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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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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.
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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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.
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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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).
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
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).
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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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
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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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
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.
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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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.
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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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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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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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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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.
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.
Annu. Rev. Ecol. Syst. 2001.32:333-365. Downloaded from
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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.
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
Allan JD. 1995. Stream Ecology: Structure and
Function of Running Waters. New York:
Chapman & Hall
Arnold CL, Boison PJ, Patton PC. 1982. Saw-
mill Brook: an example of rapid geomor-
phic change related to urbanization. J. Geol.
ArnoldCL, Gibbons CJ. 1996. Impervious sur-
face coverage: the emergence of a key envi-
ronmental indicator. Am. Planners Assoc. J.
Arthington AH. 1985. The biological resources
of urban creeks. Aust. Soc. Limnol. Bull.
Annu. Rev. Ecol. Syst. 2001.32:333-365. Downloaded from
by University of Rhode Island on 04/14/08. For personal use only.
13 Oct 2001 13:54 AR AR142-12.tex AR142-12.SGM ARv2(2001/05/10) P1: GSR
Baer KE, Pringle CM. 2000. Special problems
of urban river conservation: the encroaching
megalopolis.InGlobal PerspectivesonRiver
Conservation: Science,Policy,and Practice,
ed. PJ Boon, BR Davies, GE Petts, pp. 385–
402. New York: Wiley
Ball RC, Kevern NR, Haines TA. 1973. An
ecological evaluation of stream eutrophica-
tion.Tech. Rep. 36, Inst. Water Res., Mich.
State Univ., E. Lansing
Barringer TH, Reiser RG, Price CV. 1994. Po-
tential effects of development on flow char-
acteristics of two New Jersey streams. Water
Resour. Bull. 30:283–95
Benke AC, Wallace JB. 1997. Trophic basis of
production among riverine caddisflies: im-
plications for food web analysis. Ecology
BenkeAC, WillekeGE, ParrishFK, StitesDL.
1981. Effects of urbanization on stream
ecosystems.Environ. Resour. Ctr. Rep. 07–
81, Georgia Inst. Tech., Atlanta
Bennett EM, Reed-Andersen T, Hauser JN,
GabrielJR, Carpenter SR. 1999. A phospho-
rus budget for the Lake Mendota watershed.
Ecosystems 2:69–75
Black RW, Haggland AL, Voss FD. 2000. Pre-
dicting the probability of detecting organo-
chlorine pesticides and polychlorinated
biphenyls in stream systems on the basis of
land use in the Pacific Northwest, USA. En-
viron. Toxicol. Chem. 19:1044–54
BoetP, BelliardJ, Berrebi-dit-Thomas R, Tales
E. 1999. Multiple human impacts by the City
of Paris on fish communities in the Seine
River basin, France. Hydrobiologia 410:59–
BolandP, HanhammerS.1999. Ecosystem ser-
vicesin urban areas. Ecol. Econ. 29:293–301
Booth DB. 1990. Stream channel incision fol-
lowing drainage-basin urbanization. Water
Resour. Bull. 26:407–17
Booth DB, Jackson CR. 1997. Urbanization
of aquatic systems: degradation thresholds,
stormwater detection, and the limits of miti-
gation.J.Am. WaterResour. Assoc. 33:1077–
Borchardt D, Statzner B. 1990. Ecological im-
pact of urban stormwater runoff studied in
experimental flumes: population loss by drift
and availability of refugial space. Aquat. Sci.
BrionN, BillenG.2000.Wastewateras asource
of nitrifying bacteria in river systems: the
case of the River Seine downstream from
Paris. Water Resour. Res. 34:3213–21
Bryan EH. 1972. Quality of stormwater
drainage from urban land. Water Resour.
Bull. 8:578–88
Bryan EH. 1974. Concentrations of lead in ur-
ban stormwater. J. Water Pollut. Control Fed.
Burkholder JM. 1996. Interactions of benthic
algae with their substrata. In Algal Ecol-
ogy: Freshwater Benthic Ecosystems, ed. RJ
Stevenson,MLBothwell,RLLowe,pp. 253–
97. San Diego, CA: Academic. 753 pp.
Butler D, Davies JW. 2000. Urban Drainage.
Campbell IC. 1978. Biological investigation of
an organically polluted urban stream in Vic-
toria. Aust. J. Mar. Freshw. Res. 29:275–
Center for Watershed Protection. 2000. Urban
Stream Restoration Practices: An Initial As-
sessment.Ellicott City,MD:Cent. Watershed
Charbonneau R, Resh VH. 1992. Strawberry
Creekon the Universityof California, Berke-
ley campus—a case-history of urban stream
restoration. Aquat. Conserv. Mar. Freshw.
Ecosyst. 2:293–307
Chessman BC, Growns I, Currey J, Plunkett-
Cole N. 1999. Predicting diatom communi-
ties at the genus level for the rapid biological
assessment of rivers. Freshw. Biol. 41:317–
Chessman BC, Hutton PE, Burch JM. 1992.
Limiting nutrients for periphyton growth in
sub-alpine, forest, agricultural, and urban
streams. Freshw. Biol. 28:349–61
Chevreuiel M, Garmouma M, Fauchon N.
1999. Variability of herbicides (triazines,
phenylureas) and tentative mass balance as a
function of stream order in the River Marne
basin (France). Hydrobiologia 410:349–55
Annu. Rev. Ecol. Syst. 2001.32:333-365. Downloaded from
by University of Rhode Island on 04/14/08. For personal use only.
13 Oct 2001 13:54 AR AR142-12.tex AR142-12.SGM ARv2(2001/05/10) P1: GSR
Collier KJ. 1995. Environmental factors affect-
ing the taxonomic composition of aquatic
macro-invertebrate communities in lowland
waterways of Northland, New Zealand. N. Z.
J. Mar. Freshw. Res. 29:453–65
Collier KJ, Winterbourn MJ. 1986. Process-
ing of willow leaves in two suburban streams
in Christchurch, New Zealand. N. Z. J. Mar.
Freshw. Res. 20:575–82
Coupe RH, Manning MA, Foreman WT,
Goolsby DA, Majewski MS. 2000. Occur-
rence of pesticides in rain and air in ur-
ban and agricultural areas of Mississippi,
April-September 1995. Sci. Total Environ.
Daniels WM, House WA, Rae JE, Parker A.
2000. The distribution of micro-organic con-
taminants in river bed-sediment cores. Sci.
Total Environ. 253:81–92
Danielson MN, Keles R. 1985. The Politics
of Rapid Urbanization: Government and
Growth in Modern Turkey. New York:
Holmes & Meier
Davies PE, Cook SJ, Barton JL. 1994. Tri-
azine herbicide contamination of Tasmanian
streams: sources, concentrations, and effects
on biota. Aust. J. Mar. Freshw. Res. 45:209–
Davis JB, George JJ. 1987. Benthic inver-
tebrates as indicators of urban and motor-
way discharges. Sci. Total Environ. 59:291–
Dennehy KF, Litke DW, Tate CM, Qi SL,
McMahon PB, et al. 1998. Water quality
in the South Platte River Basin,Colorado,
Nebraska,and Wyoming,1992–1995. USGS
Circular 1167
DeVivo JC. 1995. Impact of introduced red
shiners (Cyprinella lutrensis) on stream
fishes near Atlanta, Georgia. In Proc. 1995
Georgia Water Resources Conf., ed. K
Gov., Univ. Georgia
Dickman M, Rygiel G. 1998. Municipal land-
fill impacts on a natural stream located in an
urban wetland in regional Niagara, Ontario.
Can. Field Nat. 112:619–30
Dodds WK, Welch EB. 2000. Establishing nu-
trient criteria in streams. J. N. Am. Benthol.
Soc. 19:186–96
Doeg TJ, Milledge GA. 1991. Effect of exper-
imentally increasing concentrations of sus-
pended sediment on macroinvertebrate drift.
Aust. J. Mar. Freshw. Res. 42:519–26
Douglas I. 1974. The impact of urbanization on
riversystems.InProc.Int. Geogr.UnionReg.
Conf., pp. 307–17. N. Z. Geograph. Soc.
DudaAM, LenatDR, Penrose DL. 1982. Water
quality in urban streams—what we can ex-
pect. J. Water Pollut. Control Fed. 54:1139–
Dunne T, Leopold LB. 1978. Water in Environ-
mental Planning. New York: Freeman. 818
Eganhouse RP, Simoneit BRT, Kaplan IR.
1981. Extractable organic matter in urban
stormwater runoff. 2. Molecular character-
ization. Environ. Sci. Technol. 15:315–26
Ellis JB, Marsalek J. 1996. Overview of urban
drainage: environmental impacts and con-
cerns, means of mitigation and implemen-
tation policies. J. Hydraulic Res. 34:723–31
Espey WH Jr, Morgan CW, Masch FD. 1965.
A study of some effects of urbanization on
storm runoff from a small watershed.Tech.
Rep. 44D 07–6501 CRWR-2, Ctr. for Res. in
Water Resour., Univ. Texas, Austin
Faulkner H, Edmonds-Brown V, Green A.
2000. Problems of quality designation in dif-
fusely populated urban streams—the case
of Pymme’s Brook, North London. Environ.
Pollut. 109:91–107
Finkenbine JK, Atwater DS, Mavinic DS.
2000. Stream health after urbanization. J.
Am. Water Resour. Assoc. 36:1149–60
Folke C, Jansson A, Larsson J, Costanza
R. 1997. Ecosystem appropriation by cities.
Ambio 26:167–72
Forman RTT, Alexander LE. 1998. Roads and
their major ecological effects. Annu. Rev.
Ecol. Syst. 29:207–31
FosterGD, Roberts EC Jr, GruessnerB, Velin-
sky DJ. 2000. Hydrogeochemistry and trans-
port of organic contaminants in an urban
watershed of Chesapeake Bay (USA). Appl.
Geochem. 15:901–16
Annu. Rev. Ecol. Syst. 2001.32:333-365. Downloaded from
by University of Rhode Island on 04/14/08. For personal use only.
13 Oct 2001 13:54 AR AR142-12.tex AR142-12.SGM ARv2(2001/05/10) P1: GSR
Frankie GW, Kohler CS, eds. 1983. Urban
Entomology: Interdisciplinary Perspectives.
New York: Praeger
Frick EA, Hippe DJ, Buell GR, Couch CA,
Hopkins EE, et al. 1998. Water quality
in the Appalachicola-Chattahoochee-Flint
1992–1995.USGS Circular 1164
Fusillo TV, Nieswand GH, Shelton TB. 1977.
Sediment yields in a small watershed under
suburban development. In Proc. Int. Symp.
Urban Hydrology,Hydraulics,and Sediment
Control. Lexington: Univ. Kentucky
Gafny S, Goren M, Gasith A. 2000. Habitat
condition and fish assemblage structure in
a coastal Mediterranean stream (Yargon,
Israel) receiving domestic effluent. Hydrobi-
ologia 422/423:319–30
Galli FJ. 1991. Thermal Impacts Associated
with Urbanization and Stormwater Manage-
ment Best Management Practices. Washing-
ton, DC: Washington Council of Govern-
Garie HL, McIntosh A. 1986. Distribution of
benthic macroinvertebrates in a stream ex-
posed to urban runoff. Water Resour. Bull.
Gerritse RG, Adeney JA, Dommock GM,
Oliver YM. 1995. Retention of nitrate and
phosphate in soils of the Darling Plateau in
Western Australia: implications for domestic
Gibson CJ, Stadterman KL, States S, Sykora
J.1998. Combined sewer overflows: a source
of Cryptosporidium and Giardia?Water Sci.
Technol. 38:67–72
Goni-Urriza M, Capdepuy M, Arpin C, Ray-
mond N, Caumette P, et al. 2000. Impact of
an urban effluent on antibiotic resistance of
riverine Enterobacteriaceae and Aeromonas
spp. Appl. Environ. Microbiol. 66:125–
Gonzales AE, Rodriguez MT, Sanchez JCJ,
Espinoza AJF, de la Rosa FJB. 2000. As-
sessment of metals in sediments in a trib-
utary of Guadalquiver River (Spain): heavy
metalpartitioningandrelation between water
and sediment system. Water Air Soil Pollut.
Graf WL. 1977. Network characteristics in
suburbanizing streams. Water Resour. Res.
GregoryKJ, DavisRJ, Downs PW.1992. Iden-
tification of river channel change due to ur-
banization. Appl. Geogr. 12:299–318
Gregory KJ, Gurnell AM, Hill CT, Tooth S.
1994. Stability of the pool-riffle sequence in
changing river channels. Regul. Rivers: Res.
Manage. 9:35–43
Grimm NB, Grove MJ, Pickett STA, Redman
CL. 2000. Integrated approaches to long-
term studies of urban ecological systems.
Bioscience 50:571–84
Gundacker C. 2000. Comparison of heavy
metal bioaccumulation in freshwater mol-
luscs of urban river habitats in Vienna. Env-
iron. Pollut. 110:61–71
Hachmoller B, Matthews RA, Brakke DF.
1991. Effects of riparian community struc-
ture, sediment size, and water quality on the
macroinvertebrate communties in a small,
suburban stream. Northwest Sci. 65:125–
Halling-Sorensen B, Nielsen SN, Lanzky PF,
IngerslevF, Holten-LutzhoftHC,etal. 1998.
Occurrence, fate, and effects of pharmaceu-
tical substances on the environment—a re-
view. Chemosphere 36:357–93
Hammer TR. 1972. Stream channel enlarge-
ment due to urbanization. Water Resour. Res.
Harkness N. 1982. The River Tame—a short
history of water pollution and control within
an industrial river basin. Water Sci. Technol.
Hart DD, Finelli CM. 1999. Physical-biolo-
gical coupling in streams: the pervasive ef-
fects of flow on benthic organisms. Annu.
Rev. Ecol. Syst. 30:363–95
Heaney JP, Huber WC. 1984. Nationwide as-
sessment of urban runoff on receiving water
quality. Water Resour. Bull. 20:35–42
Herlihy AT, Stoddard JL, Johnson CB. 1998.
The relationship between stream chem-
istry and watershed land cover data in the
Annu. Rev. Ecol. Syst. 2001.32:333-365. Downloaded from
by University of Rhode Island on 04/14/08. For personal use only.
13 Oct 2001 13:54 AR AR142-12.tex AR142-12.SGM ARv2(2001/05/10) P1: GSR
Mid-Atlantic region, US. Water Air Soil Pol-
lut. 105:377–86
Hill AR. 1979. Denitrification in the nitrogen
budget of a river ecosystem. Nature 281:291
Hirsch RM, Walker JF, Day JC, Kallio R.
1990. The influence of man on hydrologic
systems. In Surface Water Hydrology (The
Geology of America, Vol. O-1), ed. MG Wol-
man, HC Riggs, pp. 329–59. Boulder, CO:
Geol. Soc. Am.
Hoare RA. 1984. Nitrogen and phosphorus in
Rotorua urban streams. N. Z. J. Mar. Freshw.
Res. 18:451–54
Hoffman EJ, Latimer JS, Mills GL, Quinn
JG. 1982. Petroleum hydrocarbons in urban
runoff from a commercial land use area. J.
Water. Pollut. Control Fed. 54:1517–25
Hoffman RS, Capel PD, Larson SJ. 2000.
Comparison of pesticides in eight U.S. urban
streams. Env. Toxicol. Chem. 19:2249–58
Hogg ID, Norris RH. 1991. Effects of runoff
fromlandclearing and urbandevelopmenton
the distribution and abundance of macroin-
vertebrates in pool areas of a river. Aust. J.
Mar. Freshw. Res. 42:507–18
Horner RR, Booth DB, Azous A, May CW.
1997. Watershed determinants of ecosystem
functioning. In Effects of Watershed Devel-
opmentand Management on AquaticEcosys-
tems.ed. C Roessner, pp. 251–74. New York:
Am. Soc. Civil Eng.
HorowitzAJ, MeybeckM, IdlafkihZ, Biger E.
1999. Variations in trace element geochem-
istry in the Seine River Basin based on flood-
plain deposits and bed sediments. Hydrol.
Process. 13:1329–40
House MA, Ellise JB, Herricks EE, Huitved-
Jacobsen T, Seager J, et al. 1993. Urban
drainage—impacts on receiving water qual-
ity. Water Sci. Technol. 27:117–58
Howard KWF, Haynes J. 1993. Urban geol-
ogy. 3. Groundwater contamination due to
road deicing chemicals—salt balance impli-
cations. Geosci. Can. 20:1–8
Hunter JV, Sabatino T, Gomperts R, Macken-
zieMJ. 1979. Contribution ofurban runoffto
hydrocarbon pollution. J. Water Pollut. Con-
trol Fed. 51:2129–38
HynesHBN.1960. The Biologyof PollutedWa-
ters. Liverpool, UK: Liverpool Univ. Press
Jancarkova I, Larsen TA, Gujer W. 1997. Dis-
tribution of nitrifying bacteria in a shallow
stream. Water Sci. Technol. 36:161–66
Johnson B, Tuomori D, Sinha R. 1999. Im-
pacts of on-site sewage systems and illicit
discharges on the Rouge River. In Proc.
Natl. Conf. Retrofit Opportunities for Water
Resour. Prot. Urban Environ., pp. 132–35.
EPA/625/R-99/002. Washington, DC: EPA
Jones JB, Mulholland PJ. 2000. Streams and
Ground Waters. San Diego, CA: Academic
JonesRC, ClarkCC.1987. Impact of watershed
urbanization on stream insect communities.
Water Resour. Bull. 23:1047–55
Keefer TN, Simons RK, McQuivey RS. 1979.
Dissolved oxygen impact from urban storm
runoff. EPA-600/2–79–156. Washington,
Kennen JG. 1999. Relation of macroinverte-
brate community impairment to catchment
characteristics in New Jersey streams. J. Am.
Water Resour. Assoc. 35:939–55
KhamerM, BouyaD, Ronneau C. 2000. Metal-
lic and organic pollutants associated with ur-
ban wastewater in the waters and sediments
ofa Moroccanriver.Water Qual. Res. J. Can.
Klein RD. 1979. Urbanization and stream qual-
ity impairment. Water Resour. Bull. 15:948–
Kolkwitz R, Marsson M. 1908. Okologie der
pflanzlichen saprobien. Ber. Deutsche Bot.
Ges. 26a:505–19
Krug A. 1993. Drainage history and land use
pattern of a Swedish river system—their
importance for understanding nitrogen and
phosphorusload. Hydrobiologia251:285–96
Latimer JS, Quinn JG. 1998. Aliphatic petro-
leum and biogenic hydrocarbons entering
Narragansett Bay from tributaries under dry
weather conditions. Estuaries 21:91–107
LaValle PD. 1975. Domestic sources of stream
phosphates in urban streams. Water Res.
Leblanc RT, Brown RD, Fitzgibbon JE. 1997.
Modeling the effects of land use change
Annu. Rev. Ecol. Syst. 2001.32:333-365. Downloaded from
by University of Rhode Island on 04/14/08. For personal use only.
13 Oct 2001 13:54 AR AR142-12.tex AR142-12.SGM ARv2(2001/05/10) P1: GSR
on water temperature in unregulated urban
streams. J. Environ. Manage. 49:445–69
Lenat DR, Crawford JK. 1994. Effects of land
useonwaterquality and aquatic biotaofthree
North Carolina Piedmont streams. Hydrobi-
ologia 294:185–99
Leopold LB. 1968. Hydrology for Urban Land
Planning—A Guidebook on the Hydrologic
Effects of Urban Land Use.USGS Circular
Leopold LB. 1973. River channel change with
time—an example. Bull. Geol. Soc. Am.
LeVeenEP, WilleyWRZ.1983. A politicaleco-
nomic analysis of urban pest management.
See Frankie & Kohler 1983, pp. 19–40
Limburg KE, Schmidt RE. 1990. Patterns of
fish spawning in Hudson River tributaries:
response to an urban gradient? Ecology 71:
Litke DW. 1999. Review of phosphorus control
measures in the United States and their ef-
fects on water quality. USGS Water Resourc.
Invest. Rep. 99–4007
Mackenzie MJ, Hunter JV. 1979. Sources
and fates of aromatic compounds in urban
stormwater runoff. Environ. Sci. Technol.
MartiE, Aumatell J, Gode L, Poch M, Sabater
F. 2001. Effects of wastewater treatment
plant inputs on stream nutrient retention.
Water Resour. Res. In press
Marti E, Sabater F. 1996. High variability
in temporal and spatial nutrient retention in
Mediterranean streams. Ecology 77:854–69
Mason RP, Sullivan KA. 1998. Mercury and
methyl-mercury transport through an urban
watershed. Water Res. 32:321–30
McArthur JV, Tuckfield RC. 2000. Spatial pat-
terns in antibiotic resistance among stream
bacteria: effects of industrial pollution. Appl.
Environ. Microbiol. 66:3722–26
McConnell JB. 1980. Impact of urban storm
runoff on stream quality near Atlanta,Geor-
gia.EPA-600/2–80–094. Washington, DC:
McDonnell MJ, Pickett STA. 1990. Ecosystem
structure and function along urban-rural gra-
dients: an unexploited opportunity for ecol-
ogy. Ecology 71:1232–37
McMahon G, Cuffney TF. 2000. Quantifying
urban intensity in drainage basins for assess-
ing stream ecological conditions. J. Am. Wa-
ter Resour. Assoc. 36:1247–62
Meals DW, Levine SN, Wang D, Hoffmann
JP, CassellEA, et al. 1999.Retention ofspike
additions of soluble phosphorus in a north-
ern eutrophic stream. J. N. Am. Benthol. Soc.
MedeirosC, Leblanc R, Coler RA. 1983. An in
situ assessment of the acute toxicity of urban
runoff to benthic macroinvertebrates. Envi-
ron. Toxicol. Chem. 2:119–26
Meybeck M. 1998. Man and river interface:
multiple impacts on water and particulates
chemistry illustrated in the Seine River
Basin. Hydrobiologia 373/374:1–20
Meyer JL, McDowell WH, Bott TL, Elwood
JW, Ishizaki C, et al. 1988. Elemental dy-
namics in streams. J. N. Am. Benthol. Soc.
Meyer JL, Wallace JB. 2001. Lost linkages in
lotic ecology: rediscovering small streams.
In Ecology: Achievement and Challenge. ed.
M Press, N Huntly, S Levin, pp. 295–317.
Boston: Blackwell Sci. In press
MielkeHW, GonzalesCR, Smith MK, Mielke
PW. 2000. Quantities and associations of
lead, zinc, cadmium, manganese, chromium,
nickel, vanadium, and copper in fresh Mis-
sissippi Delta alluvium and New Orleans al-
luvial soils. Sci. Total Environ. 246:249–59
Minshall GW. 1988. Stream ecosystem theory:
a global perspective. J. N. Am. Benthol. Soc.
Moring JB, Rose DR. 1997. Occurrence and
concentrations of polycyclic aromatic hy-
drocarbons in semipermeable membrane de-
vices and clams in three urban streams of the
Dallas-FortWorthMetropolitanArea, Texas.
Chemosphere 34:551–66
Morisawa M, LaFlure E. 1979. Hydraulic ge-
ometry, stream equilibrium, and urbaniza-
tion. In Adjustments of the Fluvial System,
ed. DD Rhodes, GP Williams, pp. 333–50.
Dubuque, IA: Kendall-Hunt
Annu. Rev. Ecol. Syst. 2001.32:333-365. Downloaded from
by University of Rhode Island on 04/14/08. For personal use only.
13 Oct 2001 13:54 AR AR142-12.tex AR142-12.SGM ARv2(2001/05/10) P1: GSR
MulhollandPJ, Lenat DR. 1992. Streams of the
southeastern Piedmont, Atlantic Drainage.
In Biodiversity of the Southeastern United
States—Aquatic Communities, ed. CT Hack-
ney, SM Adams, WA Martin, pp. 193–232.
New York: Wiley
Muschak W. 1990. Pollution of street runoff by
traffic and local conditions. Sci. Total Envi-
ron. 93:419–31
Nagumo T, Hatano R. 2000. Impact of nitro-
gen cycling associated with production and
consumptionof foodon nitrogenpollution of
stream water. Soil Sci. Plant Nutr. 46:325–42
Neal C, Robson AJ. 2000. A summary of river
water quality data collected within the Land-
Ocean Interaction Study: core data for east-
ern UK rivers draining to the North Sea. Sci.
Total Environ. 251/252:585–665
Neal C, Robson AJ, Jeffery HA, Harrow
ML, NealM,et al. 1997.Trace element inter-
relationships for hydrological and chemical
controls. Sci. Total Environ. 194:321–43
Neller RJ. 1988. A comparison of channel ero-
sion in small urban and rural catchments,
Armidale, New South Wales. Earth Surf.
Process. 13:1–7
Nix PG, Daykin MM, Vilkas KL. 1994. Fecal
pollution events reconstructed and sources
identified using a sediment bag grid. Water
Environ. Res. 66:814–18
OlguinHF, SalibianA, Puig A. 2000. Compar-
ative sensitivity of Scenedesmus acutus and
Chlorella pyrenoidosa as sentinel organisms
for aquatic ecotoxicity assessments: studies
on a highly polluted urban river. Environ.
Toxicol. 15:14–22
OmernikJM.1976.The influence of land useon
stream nutrient levels. EPA-600/2–76–014.
Washington, DC: EPA
Ometo JPHB, Martinelli LA, Ballester MV,
Gessner A, Krusche A, et al. 2000. Effects
of land use on water chemistry and macroin-
vertebrates in two streams of the Piracicaba
River Basin, southeast Brazil. Freshw. Biol.
OnoratoD, Angus RA, Marion KR. 2000. His-
torical changes in the ichthyofaunal assem-
blagesoftheUpperCahaba RiverinAlabama
associatedwith extensiveurban development
in the watershed. J. Freshw. Ecol. 15:47–63
Osborne LL, Wiley MJ. 1988. Empirical rela-
tionships between land use/cover and stream
water quality in an agricultural watershed. J.
Environ. Manage. 26:9–27
Patrick R. 1973. Use of algae, especially di-
atoms, in the assessment of water quality.
In Biological Methods for the Assessment
of Water Quality, ed. J Cairns, KL Dickson,
pp. 76–95. Philadelphia: Am. Soc. Testing &
Paul MJ. 1999. Stream ecosystem function
along a land use gradient. PhD thesis, Univ.
Georgia, Athens
Pedersen ER, Perkins MA. 1986. The use of
benthic macroinvertebrate data for evaluat-
ing impacts of urban runoff. Hydrobiologia
Pereira WE, Domagalski JL, Hostettler FD,
Brown LR, Rapp JB. 1996. Occurrence and
accumulation of pesticides and organic con-
taminants in river sediment, water, and clam
tissue from the San Joaquin River and trib-
utaries, California. Environ. Toxicol. Chem.
Pesacreta GJ. 1997. Response of the stonefly
Pteronarcys dorsata in enclosures from an
urban North Carolina stream. Bull. Environ.
Contam. Toxicol. 59:948–55
PickettSTA, CadenassoML, GroveJM, Nilon
CH, Pouyat RV, et al. 2001. Urban ecolog-
ical systems: linking terrestrial ecological,
physical,and socio-economic componentsof
metropolitan areas. Annu. Rev. Ecol. Syst.
Pizzuto JE, Hession WC, McBride M. 2000.
Comparing gravel-bed rivers in paired urban
and rural catchments of southeastern Penn-
sylvania. Geology 28:79–82
Pluhowski EJ. 1970. Urbanization and its ef-
fect on the temperature of streams in Long
Island,New York.USGS Prof. Paper 627–D
Pollock JB, Meyer JL. 2001. Phosphorus as-
similationbelow a point sourcein Big Creek.
In Proc. 2001 Georgia Water Resour. Conf.,
ed. KJ Hatcher, pp. 509–9. Athens: Univ.
Annu. Rev. Ecol. Syst. 2001.32:333-365. Downloaded from
by University of Rhode Island on 04/14/08. For personal use only.
13 Oct 2001 13:54 AR AR142-12.tex AR142-12.SGM ARv2(2001/05/10) P1: GSR
Porcella DB, Sorensen DL. 1980. Character-
istics of non-point source urban runoff and
its effects on stream ecosystems.EPA-600/3–
80–032. Washington, DC: EPA
Pratt JM, Coler RA, Godfrey PJ. 1981. Eco-
logical effects of urban stormwater runoff
on benthic macroinvertebrates inhabiting the
Green River, Massachusetts. Hydrobiologia
Ragan RM, Dietmann AJ. 1976. Characteris-
tics of urban runoff in the Maryland suburbs
of Washington,DC. College Park, MD: Wa-
ter Resourc. Res. Cent., Univ. Maryland
RauchS, Morrison GM. 1999. Platinum uptake
bythe freshwater isopodAsellus aquaticusin
urban rivers. Sci. Total Environ. 235:261–68
Resh VH, Grodhaus G. 1983. Aquatic insects
inurbanenvironments.See Frankie &Kohler
1983, pp. 247–76
ReshVH, Rosenberg DM, eds.1984. TheEcol-
ogy of Aquatic Insects. New York: Praeger
Rhoads BL, Cahill RA. 1999. Geomorpholog-
ical assessment of sediment contamination
in an urban stream system. Appl. Geochem.
Richards C, Host G. 1994. Examining land use
influences on stream habitats and macroin-
vertebrates: a GIS approach. Water Resour.
Bull. 30:729–38
Riley AC. 1998. Restoring Streams in Cities: A
Guide for Planners,Policymakers,and Citi-
zens. Washington, DC: Island Press
Roberts CR. 1989. Flood frequency and urban-
induced change: some British examples.
In Floods: Hydrological,Sedimentological,
and Geomorphological Implications, ed. K
Beven, P Carling, pp. 57–82. New York:
RobinsonAM. 1976. Effects of urbanization on
stream channel morphology. In Proc. Natl.
Symp. Urban Hydrology,Hydraulics,and
Sediment Control. Univ. Ky. Coll. Eng. Publ.
III, Lexington
Roth NE, Allan JD, Erickson DL. 1996. Land-
scape influences on stream biotic integrity
assessed at multiple spatial scales. Landsc.
Ecol. 11:141–56
Sartor JD, Boyd GB, Agardy FJ. 1974. Water
pollution aspects of street surface contami-
nants. J. Water Pollut. Control Fed. 46:458–
Schueler TR. 1994a. The importance of imper-
viousness. Watershed Prot. Tech. 1:100–11
Schueler TR. 1994b. Minimizing the impact
of golf courses on streams. Watershed Prot.
Tech. 1:73–75
Schueler TR, Galli J. 1992. Environmental
impacts of stormwater ponds. In Watershed
Restoration Source Book. ed. P Kumble, T
Schueler, Washington, DC: Metropol. Wash.
Counc. Gov.
ScottJB, StewardCR, Stober QJ. 1986.Effects
of urban development on fish population dy-
namics in Kelsey Creek, Washington. Trans.
Am. Fish. Soc. 115:555–67
Seaburn GE. 1969. Effects of urban develop-
menton direct runoff to EastMeadow Brook,
NassauCounty,New York.USGSProf.Paper
Seager J, Abrahams RG. 1990. The impact
of storm sewage discharges on the ecology
of a small urban river. Water Sci. Technol.
Shutes RBE. 1984. The influence of surface
runoff on the macro-invertebrate fauna of an
urban stream. Sci. Total Environ. 33:271–82
Sloane-Richey JS, Perkins MA, Malueg KW.
1981. The effects of urbanization and storm-
water runoff on the food quality in two
salmonid streams. Verh. Int. Ver. Theor. Ang.
Limnol. 21:812–18
Smart MM, Jones JR, Sebaugh JL. 1985.
Stream-watershed relations in the Missouri
Ozark Plateau Province. J. Environ. Qual.
Steedman RJ. 1988. Modification and assess-
ment of an index of biotic integrity to quan-
tify stream quality in southern Ontario. Can.
J. Fish. Aquat. Sci. 45:492–501
Stream Solute Workshop. 1990. Concepts and
methods for assessing solute dynamics in
stream ecosystems. J. N. Am. Benthol. Soc.
Suren AM. 2000. Effects of urbanisation. In
New Zealand Stream Invertebrates: Ecology
and Implications for Management. ed. KJ
Annu. Rev. Ecol. Syst. 2001.32:333-365. Downloaded from
by University of Rhode Island on 04/14/08. For personal use only.
13 Oct 2001 13:54 AR AR142-12.tex AR142-12.SGM ARv2(2001/05/10) P1: GSR
ton: N.Z. Limnol. Soc.
Sweeney BW. 1984. Factors influencing life
history patterns of aquatic insects. See Resh
& Rosenberg 1984, pp. 56–100
TadaF, SuzukiS.1982. Adsorption anddesorp-
tion of heavy metals in bottom mud of urban
rivers. Water Res. 16:1489–94
Tate CM, Heiny JS. 1995. The ordination
of benthic invertebrate communities in the
South Platte River Basin in relation to envi-
ronmental factors. Freshw. Biol. 33:439–54
Taylor BR, Roff JC. 1986. Long-term effects
of highway construction on the ecology of
a Southern Ontario stream. Environ. Pollut.
Ser. A 40:317–44
Thorne RSJ, Williams WP, Gordon C. 2000.
The macroinvertebrates of a polluted stream
in Ghana. J. Freshw. Ecol. 15:209–17
Trimble SJ. 1997. Contribution of stream chan-
nel erosion to sediment yield from an urban-
izing watershed. Science 278:1442–44
US Census Bureau. 2001. http://www.census.
UN Population Division. 1997. Urban and Ru-
ral Areas, 1950–2030 (The 1996 Revision).
New York: United Nations
US Environ. Prot. Agency (USEPA). 2000. The
quality of our nation’s waters.EPA 841–S-
US Environ. Prot. Agency (USEPA). 2001.
US Geol. Surv. (USGS). 1999. The quality
of our nation’s waters—nutrients and pes-
ticides.USGS Circular 1225
US Geol. Surv. (USGS). 2001. http://water.
Wang L, Lyons J, Kanehl P, Bannerman R,
EmmonsE. 2000.R. Watershedurbanization
and changes in fish communities in south-
eastern Wisconsin streams. J. Am. Water Re-
sour. Assoc. 36:1173–89
Wang L, Lyons J, Kanehl P, Gatti R. 1997.
Influences of watershed land use on habi-
tat quality and biotic integrity in Wisconsin
streams. Fisheries 22:6–12
Wang X, Yin Z. 1997. Using GIS to assess
the relationship between land use and wa-
ter quality at a watershed level. Environ. Int.
Warren LA, Zimmerman AP. 1994. Suspended
particulate oxides and organic matter inter-
actions in trace metal sorption reactions in
a small urban river. Biogeochemistry 23:21–
Waschbusch RJ, Selbig WR, Bannerman RT.
1999. Sources of phosphorus in stormwater
and street dirt from two urban residential
basins in Madison,Wisconsin.USGS Water
Resour. Invest. Rep. 99–4021
Weaver LA, Garman GC. 1994. Urbanization
of a watershed and historical changes in a
streamfishassemblage. Trans.Am.Fish.Soc.
Webster JR, Benfield EF. 1986. Vascular plant
breakdown in freshwater ecosystems. Annu.
Rev. Ecol. Syst. 17:567–94
Wei C, Morrison G. 1992. Bacterial enzyme
activity and metal speciation in urban river
sediments. Hydrobiologia 235/236:597–
Weibel SR, Weidner RB, Cohen JM, Chris-
tianson AG. 1966. Pesticides and other con-
taminants in rainfall and runoff. J. Am. Water
Works Assoc. 58:1075–84
WernickBG, Cook KE, SchreierH. 1998.Land
use and streamwater nitrate-N dynamics in
an urban-rural fringe watershed. J. Am. Wa-
ter Resour. Assoc. 34:639–50
Whipple W Jr, Hunter JV. 1979. Petroleum
hydrocarbons in urban runoff. Water Resour.
Bull. 15:1096–104
Whiting ER, Clifford HF. 1983. Invertebrates
and urban runoff in a small northern stream,
Edmonton, Alberta, Canada. Hydrobiologia
Wiederholm T. 1984. Responses of aquatic in-
sects to environmental pollution. See Resh &
Rosenberg 1984, pp. 508–57
WilberWG, HunterJV.1977.Aquatic transport
of heavy metals in the urban environment.
Water Resour. Bull. 13:721–34
Wilber WG, Hunter JV. 1979. The impact
of urbanization on the distribution of heavy
metals in bottom sediments of the Saddle
River. Water Resour. Bull. 15:790–800
Annu. Rev. Ecol. Syst. 2001.32:333-365. Downloaded from
by University of Rhode Island on 04/14/08. For personal use only.
13 Oct 2001 13:54 AR AR142-12.tex AR142-12.SGM ARv2(2001/05/10) P1: GSR
Winger JG, Duthie HC. 2000. Export coeffi-
cient modeling to assess phosphorus loading
in an urban watershed. J. Am. Water Resour.
Assoc. 36:1053–61
Wolman MG. 1967. A cycle of sedimentation
and erosion in urban river channels. Geogr.
Ann. 49a:385–95
Wolter C, Minow J, Vilcinskas A, Grosch UA.
2000. Long-term effects of human influence
on fish community structure and fisheries in
Berlin waters: an urban water system. Fish.
Manage. Ecol. 7:97–104
Wood PJ, Armitage PD. 1997. Biological ef-
fects of fine sediment in the lotic environ-
ment. Environ. Manage. 21:203–17
Wright IA, Chessman BC, Fairweather PG,
Benson LJ. 1995. Measuring the impact of
sewage effluent on the macroinvertebrate
communityofan upland stream. Theeffectof
different levels of taxonomic resolution and
quantification. Aust. J. Ecol. 20:142–49
Yamamoto K, Fukushima M, Kakatani N,
KurodaK.1997. Volatileorganiccompounds
in urban rivers and their estuaries in Osaka,
Japan. Environ. Pollut. 95:135–43
Yoder CO, Miltner RJ, White D. 1999. As-
sessing the status of aquatic life designated
uses in urban and suburban watersheds. In
Proc. Natl. Conf. Retrofit Opportunities for
Water Resour. Prot. Urban Environ., pp. 16–
28. EPA/625/R-99/002
Young KD, Thackston EL. 1999. Housing den-
sity and bacterial loading in urban streams.
J. Environ. Eng. 125:1177–80
Zampella RA. 1994. Characterization of sur-
face water quality along a watershed distur-
bance gradient. Water Resour. Bull. 30:605–
Annu. Rev. Ecol. Syst. 2001.32:333-365. Downloaded from
by University of Rhode Island on 04/14/08. For personal use only.
October 9, 2001 11:23 Annual Reviews AR142-FM
Annual Review of Ecology and Systematics
Volume 32, 2001
ANECOLOGICAL PERSPECTIVE,Marcel Dicke and Paul Grostal 1
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
Anthony J. Zera and Lawrence G. Harshman 95
AREAS,S. T. A. Pickett, M. L. Cadenasso, J. M. Grove, C. H. Nilon,
R. V. Pouyat, W. C. Zipperer, and R. Costanza 127
Joanna R. Freeland, and Beth Okamura 159
APPLIED EVOLUTION,J. J. Bull and H. A. Wichman 183
WORLDWIDE,David M. Watson 219
CONSERVATION OF BIRDS,Jeffrey D. Brawn, Scott K. Robinson,
and Frank R. Thompson III 251
Eldridge S. Adams 277
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
Duncan J. Irschick and Theodore Garland Jr. 367
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RECONSTRUCTION,Peter Kershaw and Barbara Wagstaff 397
Bruce S. Rubidge and Christian A. Sidor 449
Marc Mangel, and Brent M. Haddad 481
WOLBACHIA,Lori Stevens, Rosanna Giordano, and Roberto F. Fialho 519
TROPOSPHERIC CHEMISTRY,Russell K. Monson and Elisabeth A.
Holland 547
Subject Index 577
Cumulative Index of Contributing Authors, Volumes 28–32 605
Cumulative Index of Chapter Titles, Volumes 28–32 608
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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). ...
Full-text available
Water bodies in densely populated lowland areas are often impacted by multiple stressors. At these multi-stressed sites, it remains challenging to quantify the contribution of contaminated sediments. This study, therefore, aimed to elucidate the contribution of sediment contamination in 16 multi-stressed drainage ditches throughout the Netherlands. To this end an adjusted TRIAD framework was applied, where 1) contaminants and other variables in the sediment and the overlying water were measured, 2) whole-sediment laboratory bioassays were performed using larvae of the non-biting midge Chironomus riparius, and 3) the in situ benthic macroinvertebrate community composition was determined. It was hypothesized that the benthic macroinvertebrate community composition would respond to all jointly present stressors in both water and sediment, whereas the whole-sediment bioassays would only respond to the stressors present in the sediment. The benthic macroinvertebrate community composition was indeed related to multiple stressors in both water and sediment. Taxa richness was positively correlated with the presence of PO4-P in the water, macrophyte cover and some pesticides. Evenness, the number of Trichoptera families and the SPEARpesticides were positively correlated to the C:P ratios in the sediment, whilst negative correlations were observed with various contaminants in both the water and sediment. The whole-sediment bioassays with C. riparius positively related to the nutrient content of the sediment, whereas no negative relations to the sediment-associated contaminants were observed, even though the lowered SPEARpesticides index indicated contaminant effects in the field. Therefore, it was concluded that sediment contamination was identified as one of the various stressors that potentially drove the benthic macroinvertebrate community composition in the multi-stressed drainage ditches, but that nutrients may have masked the adverse effects caused by low and diverse sediment contaminants on C. riparius in the bioassays.
... 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. ...
Full-text available
The urbanization process deeply affects rivers and streams, with numerous impacts, such as the discharge of sewers, dams, and pipework, causing profound changes in the water bodies characteristics and in their biota. In this scenario, the silting of rivers suffers one of the most impactful changes, as it undergoes a reduction in the depth and width of the rivers, triggering physical and chemical changes in the water, as well as in the structure of fish population, its feeding and reproduction habitats. As a palliative measure, it is normal to carry out the desilting (dredging) of rivers, an activity that is also very impacting. Floodings are one of the main factors that demand dredging to be carried out. This review was made to analyze desilting activities, their effects on biota and migratory fish, as well as to evaluate the best management strategies and mitigation of impacts on fish population. The shifting and removal of sediment from the riverbed can cause burial and massive death of eggs and larvae, in addition to interfering in the upward and downward migration of eggs, larvae, and adults of migratory fish. In addition, breeding and feeding sites can be impacted by sediment movement, dredging, and deposition. Some actions minimize the impacts of the silting activity recovering riparian forests, inspect the use of soil on the banks, move urban settlements away, assess the dredging site, consider the spawning sites and reduce the suspension of bottom sediments, as well as choose the best equipment and time for the performance of activities. Therefore, the development of research on the effect of dredging of water bodies on fish would contribute to a better management of the activity.
... One of the most extreme agents of landscape change is urbanization. According to Paul and Meyer (2008), the increase in impervious surface cover within catchments is the most consistent and pervasive effect resulting from urban settlement. It has been demonstrated that imperviousness strongly alters the hydrology and geomorphology of streams and rivers. ...
Freshwater ecosystems are highly interactive with the processes that occur in the surrounding basin. Any alteration of either aquatic or terrestrial ecosystems can potentially impact the structure, composition and functioning of aquatic communities. The Percy-Corintos Basin (northwestern Chubut Province) is subjected to multiple land-uses including urbanization, deforestation, extensive and intensive livestock breeding, pasture conversion, and horticulture. These local processes have had profound effects on a regional scale, altering the water quality and biodiversity of the main watercourses and associated wetlands. Livestock breeding and wood collection have resulted in an important loss of forest cover in the upper Percy basin, which has in turn accelerated erosion processes, causing sedimentation at the lower section of the basin. Urbanization has resulted in strong organic pollution, habitat impoverishment, and has decreased macroinvertebrate biodiversity in Esquel Stream. In pre-urban areas, constructed wetlands for flood prevention, act as novel environments which increase spatial heterogeneity, and consequently enhance macrophyte and aquatic invertebrate diversity. While urbanization in the lower Percy basin has a moderate effect on the river, agricultural activities like confined livestock breeding and horticulture are increasingly affecting the environmental and biological quality of the Corintos River. Management and restoration actions are urgently needed in order to restore the ecosystem functioning. The present study details and discusses specific actions to conserve biodiversity and ecosystem services.
... The awareness of the impacts of the urbanization process of the hydrological processes (Deng and Wu, 2013), climate changes (Boggs and Sun, 2011;Hao et al., 2015), and on human health (Gong et al., 2012) are shown to be fundamental factors for the investigation of the impervious surface dynamics. Paul and Meyer (2008) reported an influence of the impervious increase on physical (hydrology and geomorphology), chemical, biological and ecological streams urban landscape. The IS was treated as an essential environmental key indicator (Chester &Gibbons, 2007) for population density estimation (Morton &Yuan, 2009), changes in storm runoff (Miller, et al., 2014), water balance (Strohbach, et al., 2019) and quality (Schueler, 1994), and urban heat island (Weng, et al., 2004). ...
... Arnold & Gibbons (1996) pohojnë se në një hapësirë të pyllëzuar, rreth 90% e reshjeve që bien në atë zonë mbahen në vend: 40% e tyre rikthehen nga bimët përsëri në atmosferë nëpërmjet evapotranspirimit; dhe 50% mbeten në tokën ose nëntokën e zonës; vetëm 10% e reshjeve rrjedh në përroin më të afërt. Në një zonë të zhveshur urbane 55% të reshjeve shkojnë menjëherë në kanal, evapotranspirimi zbret deri në 30%, kurse ujërat në tokë dhe nëntokë pakësohen shumë (deri në 15%), duke nxitur më tej përmbytjet, shembjet e dheut, por dhe ashpërsimin e klimës së zonës (Paul & Meyer, 2001). ...
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Në këtë punim jepen të dhëna mbi cilësinë e ajrit, zhurmat dhe gjelbërimin në një zonë urbane në qendër të Tiranës (48 ha; rreth 11,000 banorë), pjesë e njësive 9 dhe 10. Matjet janë kryer në 22 stacione në Maj 2019 (ajri) dhe Shkurt 2020 (zhurmat); kurse gjelbërimi në zonë është vlerësuar gjatë periudhës Dhjetor 2018-Gusht 2020; janë vlerësuar llojet bimore në gjithë hapësirën publike; në veçanti drurët dhe shkurret mbi 1 m janë numëruar dhe digjitalizuar në hartë; është përcaktuar mosha dhe shërbimet kryesore që ato ofrojnë për zbutjen e ndotjes; çlirimi i O2, përvetimi i CO2 dhe kursimi i energjisë kWh, janë përllogaritur në vlerë monetare nëpërmjet modelit kompiuterik “National Tree Benefit Calculator”. Të dhënat janë digjitalizuar në Sistemin e Informacionit Gjeografik (GIS). Mesatarja e 22 matjeve për gjithë parametrat e cilësisë së ajrit tejkalon normat e standardit shqiptar: PM2.5, 29 μg/m3 (16% mbi Stash); PM10, 62.2 μg/m3 (4%); CO2, 475.7 mg/l (36%); NO2 125.7 μg/m3 (214%). Në zonë u gjetën 276 lloje bimësh; 88 lloje ishin drurë dhe shkurre mbi 1 m, gjithsej 2,365 individë, të përfaqësuar kryesisht nga bliri, voshtra, panja amerikane, ilqja, rrapi, tuja etj. Mosha mesatare e tyre ishte rreth 25 vjeç dhe lartësia mesatare 7.15 m. Sipërfaqja e përgjithshme e blertë është rreth 13,500 m2 (2.81% e zonës), ose rreth 1.23 m2/banorë; krahasuar me vitin 2007 sipërfaqja e blertë në këtë zonë është zvogëluar me 8,690 m2 (39%). Edhe pse rastësor, vlerësimi ynë përbën një sinjal të fortë për gjëndjen sot në Tiranën urbane. Këshillohet shumë njohja e gjendjes nga strukturat përgjegjëse, në vazhdimësi dhe në hapësirë; por më shumë këshillohet kujdesi për hapësirat ekzistuese, gjelbërimi mundësisht me bimë vendase. Monitorimi është pjesë e projektit të BE ‘Mushkëri të gjelbra për qytetet tona’, që po zbatohet nga Co-PLAN, si mundësi monitoruese alternative.
... 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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Urbanisation alters the natural hydromorphology of streams, affecting aquatic communities and ecological quality. Increasing efforts have been put into the rehabilitation of urban streams due to their importance for urban sustainability. Despite these efforts, many projects fail to achieve the improvement of aquatic communities. This study aims to provide specific recommendations to enhance the biological rehabilitation of urban streams by reviewing: (i) the impacts of urbanisation and climate change on urban stream hydrology, (ii) the responses of invertebrate assemblages to alterations in the hydrology and morphology of streams, and (iii) the hydromorphological rehabilitation measures applied to streams and their effect on invertebrate communities. This review found that commonly employed measures of habitat heterogeneity enhancement (such as the addition of meanders, boulders, and artificial riffles) are not enough to improve invertebrate communities. On the other hand, the most effective measures are those leading to the re-establishment of natural hydrological patterns and good water quality. Ultimately, an integrated ecohydrological approach that considers the entire watershed and its interactions between ecosystems and anthropological activities is the key to managing and rehabilitating urban streams.
In Te Whanganui‐a‐Tara Wellington, Aotearoa New Zealand, 95% of the length of urban streams are piped and buried, leaving them hidden. Technocratic management regimes separate streams into discrete parts, and thus overlook important relationships that can improve or maintain the state of streams. This paper takes a hydrosocial perspective on urban streams, exploring connections between water flow, infrastructure, and social structures and institutions to consider the buried Waimapihi Stream. Drawing on interviews, it examines implications of the Waimapihi's legislative status and management for its headwaters at the Waimapihi Reserve and the stormwater piped section, and considers the potential for daylighting the piped section.
Urbanization increases runoff, sediment, and nutrient loadings downstream, causing flooding, eutrophication, and harmful algal blooms. Stormwater control measures (SCMs) are used to address these concerns and are designed based on inflow loads. Thus, estimating nutrient and sediment loads is important for meeting restoration objectives. Pollutants accumulate on surfaces during dry periods, making Event Mean Concentration (EMC) a function of antecedent dry period (ADP). An EMC results from wash-off of accumulated pollutants from catchment surface during runoff events. However, several studies found little to no correlation between constituent concentrations in stormwater and ADP. The objective of this study is to verify this finding and discover which climatological or catchment characteristics most significantly affect stormwater quality. Stormwater quality data were obtained from the National Stormwater Quality Database (NSQD), which is the largest data repository of stormwater quality data in the U.S. Bayesian Network Structure Learner (BNSL) was used to assess the relationships between catchment characteristics, climatological information, and stormwater quality for selected land uses. Given the optimal BN structure, it was determined which parameters most affect stormwater quality EMCs. The results demonstrate that both catchment and rain characteristics affected stormwater quality EMCs. Among catchment characteristics, land use (LU) was the most important factor and catchment size was the least. Precipitation depth (P) and duration (D) affected Total Phosphorus (TP), Total Nitrogen (TN), and Total Suspended Solids (TSS). This indicated that it is likely that P and D had a greater influence on stormwater quality more than ADP. P, D, and ADP affected the dissolved constituents of TN (i.e. NO2-N/NO3-N) and TP (i.e. Ortho-P). Compared to other factors (i.e. P and D), the effect of ADP on TSS was negligible. Stormwater quality EMCs related to nitrogen were not affected by catchment slope (S). However, TSS and Ortho-P were influenced by S.
Habitat alteration and destruction are primary drivers of biodiversity loss. However, the evolutionary dimensions of biodiversity loss remain largely unexplored in many systems. For example, little is known about how habitat alteration/loss can lead to phylogenetic deconstruction of ecological assemblages at the local level. That is, while species loss is evident, are some lineages favored over others? Using a long-term dataset of a globally, ecologically important guild of invertebrate consumers, stream leaf “shredders,” we created a phylogenetic tree of the taxa in the regional species pool, calculated mean phylogenetic distinctiveness for >1000 communities spanning >10 y period, and related species richness, phylogenetic diversity and distinctiveness to watershed-scale impervious cover. Using a combination of changepoint and compositional analyses, we learned that increasing impervious cover produced marked reductions in all three measures of diversity. These results aid in understanding both phylogenetic diversity and average assemblage phylogenetic distinctiveness. Our findings suggest that, not only are species lost when there is an increase in watershed urbanization, as other studies have demonstrated, but that those lost are members of more distinct lineages relative to the community as a whole.
Natural disturbances play important roles in the functioning and structure of lotic ecosystems, especially in small streams. Adaptation to natural disturbances, in the form of resilience, can be affected by anthropogenic disturbances such as urbanization and industrial zones, which in turn can limit stream biodiversity. The aim of this study was to assess the effects of runoff from urban and industrial zones on the resilience of benthic macro-invertebrate assemblages in small streams. For that, we tested the hypothesis that benthic macroinvertebrate assemblages in streams affected by urbanization and industrialization have lower resilience to natural disturbances than those in reference areas. We calculated the recovery proportions of Taxa Richness, Taxa Abundance, Resistant Taxa Richness, Resistant Taxa Abundance, Sensitive Taxa Richness and Sensitive Taxa Abundance. Recovery proportions of freshwater biodiversity were calculated as the target variable values during the dry season divided by the same variable in the previous rainy season. Taxa Richness recovery proportion and Sensitive Taxa Richness recovery proportion were significantly higher (p < 0.01) in the reference sites. Resistant Taxa Richness and Sensitive Taxa Abundance followed the same pattern but were less significant (p < 0.1). These results indicate that streams draining urban and industrial areas have significantly lower resilience to natural disturbances than their counterparts in reference areas. Our results also suggest that both landscape and local environmental conditions play important roles in maintaining naturally resilient lotic ecosystems and biodiversity in the neotropics.
Intermittent discharges of storm sewage from combined sewer overflows continue to be one of the principal causes of poor water quality in many urban rivers in the UK. Despite the persistent nature of this problem, very little attention has been given to the study of how discharges of varying magnitude, duration and frequency affect the ecological quality of receiving waters. This information is of critical importance for deriving meaningful water quality criteria for the control of intermittent pollution. This paper describes the results of a study which has been carried out on Pendle Water, a river which flows through the urban catchment of Burnley, Lancashire, UK. Both the chemical and biological quality of Pendle Water are adversely affected by storm sewage discharges during heavy rainfall events. The ecological investigation has been primarily concerned with impact of these episodic discharges on benthic invertebrate communities and physiological responses in fish. Quantitative sampling of macroinvertebrates has indicated that storm sewage discharges may have a significant impact on the structure and diversity of benthic communities in receiving waters. Physico-chemical properties of habitats appear to be altered in a way which tends to favour the proliferation of certain pollution-tolerant species and decrease the abundance of taxa intolerant of organic pollution. Insitu bioassays, including the WRc Mark III Fish Monitor, have been deployed to investigate physiological responses to storm events of different magnitude, duration and frequency. Results are discussed in relation to their application in the field validation of proposed water quality criteria for the control of intermittent pollution from combined sewer overflows.
A project investigating the dynamics of self-purification processes in a shallow stream is carried out. Effects of the concentration gradient due to the distance to the pollution source, of hydraulic conditions in the river bed and of storm floods on the distribution of nitrifying bacteria were studied with the help of laboratory and field experiments. Nitrifiers density on the surface of the stream bed increased rapidly up to a distance of 300 m from the WWTP indicating possible competition of the nitrifiers with the heterotrophic bacteria close to the WWTP. Afterwards a slight decrease in the downstream direction was observed. In vertical profiles, higher bacterial densities were found at sites with rapid infiltration of channel water to the stream bed than at sites with no exchange between channel water and stream bed water or where stream bed water exfiltrated. A major flood event scoured the nitrifiers nearly totally from the surface of the river bed. Major floods belong so to the most dominant processes controlling self-purification in shallow streams. Minor floods, however, don't scour bacteria in the depth of the stream bed that could then be important for the self-purification processes.
Padden Creek, a second-order stream in Bellingham, Washington, was studied for patterns in the macroinvertebrate community structure between upstream and downstream sites. Many taxa were more abundant at the undisturbed, forested upstream site, especially pollution intolerant mayflies, stoneflies, and caddisflies. Where the surrounding forest had been cleared, there was a decrease in the density of shredders, predators, and collector-filterers, and an increase in scrapers. Where the stream is affected by organic pollution, there were fewer representatives from pollution intolerant orders (Ephemeroptera, Plecoptera, and Trichoptera) and many more non-insect taxa (e.g., oligochaetes and gastropods). Thus, the effects of channelization, deformation, and pollution resulted in major changes in the structure of macroinvertebrate communities at downstream sites, suggesting that such riparian alterations imitate similar urbanization effects of higher-order rivers. -from Authors
The impact of wastewater on the Sebou river, one of the most important water resources in Morocco, is analyzed. A significant amount of wastewater is discharged every day in this aquatic system by the city of Fez. Heavy metals and toxic elements (Cr, Zn, Fe, Co, As, La, Sr, Ag, Hg, Sb and Rb) were identified by neutron activation in the urban wastewater and in the waters and sediments of the river both upstream and downstream from their point of origin. Results show high contamination levels. The highest levels were for Cr and Zn, while Hg, Ag and As were found only at the downstream sites where there was also a large increase in organic loadings (COD, BOD, volatile organic matter), salinity and ammonium, orthophosphates, sulfate, sodium and calcium, as well as a sharp drop in dissolved oxygen. Aside from the decline in water quality, there was a large decrease in the biotic index, from 8 upstream to 2 downstream. As a result of the water regime of this river and the climatic conditions in the region, heavy metals accumulating in the sediments could be remobilized and create the risk of large-scale contamination.
A study of 8 streams draining basins of about 1 square mile in the Piedmont province of the Baltimore, Maryland-Washington, D. C. metropolitan area indicates that urbanization, resulting in increased magnitude and frequency of flood flows and changes in watershed sediment yield, has considerable impact upon stream channel morphology. Three watersheds were rural, three urban, one partially urbanized, and one was undergoing development. Width, depth, wetted perimeter, and cross-sectional area were determined at twenty sections for each stream. Channel area of the urban streams was on the order of 2 times and width/depth ratios 1. 7 times those of the rural stream channels. The size distribution of bed material was altered by a reduction in the fractions of silt and sand as well as by an increase in the cobble fractions. Channel geometry progressively changes. It is postulated that a new equilibrium form is not achieved for at least fifteen years.