Urban Sprawl Leaves Its PAH

Signature

PETER C. VAN METRE,* , †

BARBARA J. MAHLER, † AND

EDWARD T. FURLONG ‡

U.S. Geological Survey, 8027 Exchange Drive,

Austin, Texas 78754, and U.S. Geological Survey,

P.O. Box 25046, MS 407, Denver Federal Center,

Lakewood, Colorado 80225

The increasing vehicle traffic associated with urban

sprawl in the United States is frequently linked to degradation

of air quality, but its effect on aquatic sediment is less wellrecognized. This study evaluates trends in PAHs, a

group of contaminants with multiple urban sources, in

sediment cores from 10 reservoirs and lakes in six U.S.

metropolitan areas. The watersheds chosen represent a

range in degree and age of urbanization. Concentrations of

PAHs in all 10 reservoirs and lakes increased during the

past 20-40 years. PAH contamination of the most recently

deposited sediment at all sites exceeded sedimentquality guidelines established by Environment Canada, in

some cases by several orders of magnitude. These results

add a new chapter to the story told by previous coring

studies that reported decreasing concentrations of PAHs

after reaching highs in the 1950s. Concurrent with the increase

in concentrations is a change in the assemblage of

PAHs that indicates the increasing trends are driven by

combustion sources. The increase in PAH concentrations

tracks closely with increases in automobile use, even in

watersheds that have not undergone substantial changes

in urban land-use levels since the 1970s.

Introduction

Polycyclic aromatic hydrocarbons (PAHs) represent the

largest class of suspected carcinogens (1) and can present a

threat to aquatic life (2). The presence and distribution of

PAHs in the environment are largely a product of the

incomplete combustion of petroleum, oil, coal, and wood

(3). Anthropogenic sources such as vehicles, heating and

power plants, industrial processes, and refuse and open

burning are considered to be the principal sources to the

environment (4). On the basis of 1989 data, vehicles produced

11% of PAH emissions in the United Kingdom, domestic

coal burning produced 84%, and industrial processes produced 3% (5). Several studies in the 1970s and 1980s reported

decreasing trends in PAH concentrations in the environment

on a regional scale (United States and Europe) since their

peak in the 1950s and 1960s (6-9), on the basis of data from

sediment cores from remote and urban lakes and rivers. These

reductions have been attributed to reduced use of coal for

home heating, industrial emissions controls, and increased

efficiency of power plants (7, 9-11).

While loads of PAHs from some sources may have

decreased, the changing face of the urban landscape has

resulted in an increase in another source of PAHs: vehicle

use. Growth in the use of land for residential and commercial

purposes in the United States now far outstrips growth in

population (12), a phenomenon termed urban sprawl.

Increasing sprawl has resulted in decentralized employment

and workplace facilities and greater dependence on vehicles,

as reflected in number of miles traveled per capita and per

vehicle (13, 14). What effect, if any, has this change had on

urban water bodies?

As a part of the U.S. Geological Survey National Water

Quality Assessment (NAWQA) Reconstructed Trends program, trends in PAHs were tracked over the last several

decades to the mid-to-late 1990s in sediment cores from 10

lakes and reservoirs in six U.S. metropolitan areas. This study

indicates that trends in PAH concentrations in developed

watersheds over the last three decades, rather than decreasing, are increasing and that the increases may be linked to

the increasing amount of urban sprawl and vehicle traffic in

urban and suburban areas.

Methods

Sediment cores from lakes and reservoirs can be used to

reconstruct historical trends in water quality (15, 16). Trends

are recorded for hydrophobic, persistent compounds, such

as PAHs, that bind to sediment particles. PAHs sorb to

particulates and are transported to receiving water bodies

via atmospheric deposition (17), sewage effluent (18), and

surface runoff (19). For this study, sediment cores were

collected from seven reservoirs and three lakes across the

United States (Figure 1). Land use in the watersheds of these

reservoirs and lakes is largely mixed residential and commercial with percent urban land use ranging from 25 to nearly

100% (Table 1). Three sites are in watersheds experiencing

rapid growth since the 1970s (58-122% increase in urban

land use), three sites are in watersheds experiencing moderate

growth (26-36%), and four sites are in watersheds with

relatively stable levels of urban land use (0-5%). The onset

of development ranges from the early 1900s (e.g., Lake Harriet

and Newbridge Pond) to as recently as the 1970s (e.g., Palmer

Lake).

Cores were collected from the deepest part of the lake or

in the lower part of the reservoir and sectioned into vertically

discrete subsamples for analysis of 137Cs and PAHs. 210Pb was

analyzed for selected lakes. Samples were also analyzed for

major and minor elements, chlorinated organic pesticides,

and PCBs (not presented in this paper).

* Corresponding author phone: (512)927-3506; (512)927-3590;

e-mail: pcvanmet@usgs.gov. † U.S. Geological Survey, Austin, TX.

‡ U.S. Geological Survey, Lakewood, CO.

FIGURE 1. Locations of sampling sites. Abbreviations correspond

to lake and reservoir names given in Table 1.

Environ. Sci. Technol. 2000, 34, 4064-4070

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PAHs and alkyl-substituted PAHs (alkyl-PAHs) were

extracted, isolated, and analyzed using a variation of the

procedure of Furlong et al. (23). Briefly, wet bottom sediment

was extracted overnight with dichloromethane in a Soxhlet

apparatus. The extract was reduced and filtered. An aliquot

of the extract was processed by automated gel-permeation

chromatography (GPC) using a styrene-divinyl benzene

column and a dichloromethane eluent to remove sulfur and

partially isolate the PAHs from coextracted high molecular

weight interferences such as humic substances. The PAH

fraction was reduced in volume and solvent-exchanged to

ethyl acetate in a micro-Snyder column. An aliquot of internal

standard solution (a mixture of perdeuterated PAHs) was

added to the final extract. PAHs were separated, identified,

and quantified by capillary gas chromatography (GC) coupled

to mass spectrometry (MS). The parent and alkyl-PAHs were

resolved on a fused silica capillary GC column. Selected ion

monitoring (SIM) was used to reduce chemical interferences

and improve sensitivity. Parent PAHs were identified and

quantified by comparison to authentic standards. Individual

alkyl-PAHs were quantified when authentic alkyl-substituted

standards were available. The multiple isomeric alkyl-PAHs

were quantified from SIM mass chromatograms as the sum

of all isomers at each alkylation level (C1-naphthalene, C2-

naphthalene, etc.). When authentic alkyl-substituted standards were unavailable, a parent PAH was used for quantitation. Nineteen parent PAHs, 10 specific alkyl-PAHs, and

the homologous series of alkyl-PAHs were determined for

this study. Total PAHs (∑PAH) was computed as the sum of

these, excluding perylene.

Sediment 137Cs activity profiles were measured in all 10

reservoirs and lakes (Table 2) by counting freeze-dried

sediments in fixed geometry with a high-resolution, intrinsic

germanium detector γ-spectrometer. The method is similar

to that reported by Callender and Robbins (24). Activity

concentrations of 210Pb were measured in two older reservoirs

and two lakes (Table 2) by high-precision γ-ray spectrometry.

A low-energy photon planar detector was used to detect γ-ray

emissions between 40 and 400 keV (including 210Pb), and a

high-purity germanium coaxial detector was used for γ-ray

emissions between 200 and 2000 keV (including 137Cs).

Date of deposition for sediment intervals within each core

was based on a variety of date-depth markers. Those

indicators on which dates were based and those used to

corroborate the dates assigned are listed in Table 2. The depth

of the pre-reservoir surface in cores, for example, was

matched with reservoir construction date in 5 of the 7

reservoirs. (In the other two, Newbridge Pond and Orange

Reservoir, the age of the reservoirs and the presence of DDT,

PCBs, and 137Cs immediately above the pre-reservoir surface

indicated a discontinuity.) The other major date-depth

markers used were 137Cs first occurrence (1953.0), 137Cs peak

concentration (1964.0), and sampling date. Additional and/

or corroboratory date-depth markers included unsupported 210Pb profile (Lake Harriet only), lead peak (mid-1970s (25)),

and DDT and PCB profiles (first occurrence in the 1940s and

peaks in the early and late 1960s, respectively (16)).

Cesium-137 profiles provided date markers in 8 of the 10

reservoirs and lakes. The core from White Rock Lake is a

good example: 137Cs is first detected at a depth of 58 cm,

peaks sharply at 48 cm, and then exhibits a smooth

exponential decrease to the top of the core. In addition to

providing age control, this type of profile is strong evidence

that sediments have not been disturbed by bioturbation or

other post-depositional mixing. At one site, Lake Harriet,

because the core penetrated sediment deposited long before

the release of 137Cs into the environment 210Pb was used in

addition to 137Cs to provide date information for the lower

TABLE 1. Selected Characteristics of Lakes Sampled

reservoir or lake

sampling

date

depth of

water (m)

lake area

(km2)

watershed

area (km2

)

% urban,

1970sa

% urban,

1990b

% urban change,

1970s-1990c

Lake Anne, VA (ANN) Jun 1996 6.5 0.13 2.3 44.8 85.7 91

Lake Ballinger, WA (BAL) Jun 1998 9.9 0.40 13.7 93.0 96.5 4

Lake Fairfax, VA (FFX) Sep 1997 3.8 0.11 8.4 58.8 92.8 58

Lake Harriet, MN (HAR) Jul 1997 21.3 1.2 6.1 79.7 79.7 0

Newbridge Pond, NY (NEW) Sep 1997 2 0.042 7.9 99.6 99.6 0

Orange Reservior, NJ (NJOR) Sep 1997 nrd 0.35 11.7 77.1 97.1 26

Lake Packanack, NJ (PAK) Sep 1997 3 0.33 4.8 73.2 93.2 27

Palmer Lake, MN (PLM) Jul 1997 1 0.13 64.9 29.6 65.6 122

Town Lake, TX (TWN) Aug 1998 8.5 0.73 404 23.8 25.1e 5

White Rock Lake, TX (WRL) Jun 1996 4.9 4.4 265 56.4 76.7 36

a Based on ref 20. b Based on ref 21. c Sites grouped by percent change in urban land use: rapid (ANN, FFX, PLM), moderate (NJOR, PAK, WRL),

and stable (BAL, HAR, NEW, TWN) ubranization. d nr, not recorded. e Based on ref 22.

TABLE 2. Sediment Accumulation Rates and Age Assignments in Core

lake

basis of

age datinga

corroberating

evidence for

age datinga

approx time

period of core

sediment

thickness

(cm)

av linear

sedimentation

rate (cm/yr)

av mass

accumulation

rate (g cm-2 yr-1

)

Lake Anne a, c, e, f d 1968-1996 17 0.60 0.18

Lake Ballinger c d, e 1947-1998 23 0.45 0.11

Lake Fairfax a, c d 1952-1997 52 1.2 0.86

Lake Harriet b, c d, e ∼1800-1997 16 0.60 0.07

Newbridge Pond c d, e 1952-1997 54 1.2 0.32

Orange Reservoir d, e e 1949-1997 32 0.66 0.24

Lake Packanack a d, f, Cub 1932-1997 42 0.64 0.11

Palmer Lake c, d e 1949-1997 38 0.78 0.41

Town Lake a, c d, e 1959-1998 110 2.7 1.95

White Rock Lake a, c d, e 1913-1996 105 1.3 1.13

a a, construction date of reservoir; b, 210Pb; c, 137Cs; d, DDT and(or) PCB profiles; e, total Pb peak; f, modeled exponential decrease in sedimentation

rate following approach of ref 24. Sampling date used at top of all cores. b Dramatic trends in copper were matched to historical copper-sulfate

treatments of the lake.

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portion of the core, following the constant delivery rate model

(26).

A date-depth marker other than 137Cs was used for 2 of

the 10 lakes (NJOR and PAK) because the 137Cs peak was not

sufficiently defined. For NJOR, a well-defined DDT peak was

evidence of undisturbed sediments, and a date of 1964 was

assigned to that depth. For PAK, dates were assigned based

on occurrence of Cu: treatment of the lake with CuSO4 began

in about 1953, which we correlated to the sharp increase in

Cu in the core at 18 cm. For all cores, other date markers

such as the DDT peak in the early to mid-1960s and the lead

peak in the mid-1970s were checked to see if their depths

within the cores were consistent with the dates assigned.

Lakes Anne and Packanack were the only two lakes for

which the assumption of constant sedimentation rate seemed

questionable. In these two lakes the initial dating scheme

placed the lead peak in the 1980s, unrealistically late. This

indicated that mass sedimentation rates must have decreased

substantially between the occurrence of the older date

markers (137Cs for Anne and Cu for Packanack) and the top

of the core. We therefore assumed that a gradual change in

sedimentation rate occurred, possibly caused by gradual

completion of urban construction in the watersheds resulting

in a corresponding reduction in erosion. We modeled this

change using an exponential function following the approach

of ref 24, where an exponential decrease in mass accumulation rate was demonstrated in many reservoirs.

Although confidence intervals cannot be assigned to the

dates corresponding to individual sample intervals in the

cores, the consistency of multiple date markers suggests that

FIGURE 2. Normalized ∑PAH and ratios of PAH assemblages in sediment cores. Trends in ∑PAH among sites are compared by normalizing

concentrations in samples from each core to the most recently deposited sample, thus showing trends relative to current conditions.

Normalized ∑PAH are shown for watersheds with recent rapid increases (a), moderate increases (b), and stable levels (c) of urbanization.

Ratios of PAH assemblages as indicators of PAH source are shown for recent rapid increases (d), moderate increases (e), and stable levels

(f) of urbanization. Abbreviations for lake and reservoir sites are as given in Table 1.

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most dates are probably within a few years of the actual

deposition dates. Pronounced peaks and systematic variations in chemical profiles indicate that contaminant trends

have not been obscured by post-depositional mixing. We

conclude therefore that these cores are recording multi-year

to decadal trends in the contaminants of interest.

Results

Modest to dramatic increases in ∑PAH concentrations are

seen in sediments deposited in all 10 watersheds over the

last 20-40 years (Figure 2a-c). In all three rapidly urbanizing

sites, ∑PAH has increased sharply from pre-development

levels to the present (Figure 2a). Recent concentrations were

1-2 orders of magnitude above pre-development concentrations at these sites. Among the moderately urbanizing

and stable sites, differences in trends reflect the age of the

onset of urbanization (Figure 2b,c). In those reservoirs in

which the onset of urbanization was largely post-1960, ∑PAH

increases steadily from the date of reservoir construction to

the present, similar to the trends seen in the rapidurbanization watersheds. In the three watersheds that

underwent pre-1960s urbanization and in which the lake or

reservoir was in place to record water-quality changes (HAR,

PAK, and NJOR), ∑PAH peaked in the 1950s, similar to trends

reported at other sites by other investigators (6-9), and then

decreased (Figure 2b,c). In each case, however, this decrease

was followed by a subsequent increase beginning in the 1960s

in PAK and NJOR and in the 1980s in HAR.

Although many characteristics of the temporal trends are

similar at these sites, the magnitude of concentrations varies

greatly (Table 3). Concentrations of PAHs are a function not

only of source strength but also of sedimentation rate and

sediment dilution and thus can vary widely from site to site.

The smallest concentrations are in White Rock Lake in Dallas,

TX, with a ∑PAH at the sediment surface of 2790 μg/kg. The

largest concentrations are in Newbridge Pond, a small

reservoir in an older residential and commercial neighborhood on Long Island in New York, with a ∑PAH in recent

sediments of 224 000 μg/kg. Intermediate concentrations,

ranging from about 10 000 to 50 000 μg/kg in surficial

sediments, occur at the other eight sites.

The assemblage of PAH compounds in the cores indicates

a general shift in PAH source over the last 40 years from

uncombusted to combusted fossil fuels, coincident with

increased concentrations and urbanization. Uncombusted

sources (e.g., oil seeps, petroleum spills) contain predominantly two- and three-ringed compounds, whereas combustion (e.g., vehicle exhaust, domestic heating with coal, forest

fires) results in predominantly four- and five-ringed species

(7, 18). Figure 2d-f shows the change over time in the ratio

of two- and three-ringed PAH compounds plus homologues

to the sum of the major “combustion” PAHs (28). The

combustion PAHs are fluoranthene, pyrene, benz[a]anthracene, chrysene, benzo[b]fluoranthene, benzo[k]fluoranthene,

benzo[e]pyrene, benzo[a]pyrene, indeno[1,2,3-cd]pyrene,

and benzo[g,h,i]perylene. A decrease in this ratio indicates

TABLE 3. Concentrations of Selected PAHs in Cores

surficial (top of core) concn range in concn in corea

lakeb benzo[a]pyrene fluoranthene pyrene total PAH benzo[a]pyrene fluoranthene pyrene total PAH

White Rock Lake 154 262 219 2,790 e3-154 e9-317 e8-262 135-3440

Town Lake 585 1320 1100 11,400 33-582 57-1,320 66-1,100 670-11 400

Lake Packanack 528 815 754 12,100 41-744 147-960 118-1,000 1610-15 900

Orange Reservoir 913 1480 1260 21 600 314-1500 522-2660 524-2370 13 000-29 100

Lake Anne 1020 4850 3410 30 300 20-1020 62-4850 54-3410 1030-30 300

Lake Fairfax 1540 3500 2680 30 800 124-1540 346-4170 278-2690 3410-30 800

Lake Harriet 1960 4210 3160 35 900 e8-3430 39-16 800 26-12 700 430-48 300

Palmer Lake 2110 5480 4380 44 000 19-2110 65-5930 46-4600 518-45 700

Lake Ballinger 2350 4120 4340 46 300 72-2910 185-7420 263-7430 4810-49 000

Newbridge Pond 10 800 27 200 22 100 224 000 597-10 800 1450-27 200 1,590-22 100 26 300-224 000

Canadian SQGsc 31.9/762 111/2355 53.0/875

a Estimated values are maked with ‘e’. b Lakes are ordered by total PAH concentrations in surficial samples. c Canadian sediment-quality guidelines;

interim sediment-quality guideline/probable effect level (27).

FIGURE 3. Comparison of percent increase in ∑PAH to percent increase in urban land use for each watershed (based on refs 21 and 22)

from 1975 to 1995.

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a shift from uncombusted to combusted fossil fuels as the

PAH source. The proportion of noncombustion sources

relative to combustion sources increased during the early

history of four watersheds (WRL, NEW, NJOR, and BAL),

FIGURE 4. Comparison of increases in vehicle miles traveled (VMT) to changes in ∑PAH. Vehicle miles traveled per year over the entire

metropolitan center (14) are shown by filled circles; ∑PAH for the lake or reservoir from the watershed within that area are shown by

shaded symbols.

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peaking from the 1950s to the 1970s; all four sites had some

degree of pre-1960s urban development. In contrast, over

the last few decades, 9 of the 10 sites show an increase in the

importance of combustion sources relative to noncombustion

sources coincident with increasing concentrations. Only PAK

shows a varying ratio with no trend.

The change in ∑PAH from the mid-1970s to the mid1990s is compared to the change in the amount of urban

land use for each watershed for the same period (Figure 3).

In 9 out of 10 cases (PAK being the exception), increase in

∑PAH outpaces increase in urbanization. At some sites (BAL,

NEW, and TWN), there was virtually no increase in the

amount of urban land use during that period, yet ∑PAH more

than doubled.

To investigate the possible effect of traffic on trends in

∑PAH, trends in vehicle use for the six metropolitan areas

in this study were plotted versus ∑PAH (Figure 4). Trends in

vehicle miles traveled (VMT) for each metropolitan area and

∑PAH for the 1970s-1990s show similar increases. This

comparison is limited in that the VMT data are for entire

metropolitan areas and are not specific to the sampled

watersheds. In most cases the watersheds contributing runoff

to the sampled lakes are much smaller than the urban area,

and changes in traffic volume over time may not necessarily

match those of the larger urban area. The possible correlation

can be better investigated by looking at two reservoir

watersheds that correspond well to the urban areas of the

VMT data: Town Lake (TWN) in Austin and White Rock Lake

(WRL) in Dallas, TX. Town Lake receives drainage from most

of urban Austin; nonurban sediments from the Colorado

River, which forms Town Lake, are trapped by a series of

upstream reservoirs. The White Rock Lake Watershed covers

264 km2 of diverse urban land use that is generally a similar

mix of residential, commercial, industrial, and transportation

land uses as greater Dallas (29). For these two sites, the slopes

of PAH concentrations and VMT versus time are very similar

(Figure 4), indicating that increases in PAHs are proportional

to increases in vehicle traffic, at least for these two locations.

Discussion

The results of this study indicate a reversal in the decreasing

trend in PAHs in older urban watersheds; a rapid increase

in PAHs in all watersheds over the last two decades, resulting

in severe degradation of sediment quality at most of these

sites; and a possible relation between vehicle traffic and recent

deterioration of sediment quality with respect to PAHs in

receiving water bodies in urban areas of all ages.

The decreasing trends in PAHs reported in studies carried

out in the 1970s and 1980s have been largely attributed to

the transition from home heating with coal to the use of oil

and natural gas and increased efficiency of centralized power

plants (6-9). A possible reversal of the decreasing trend is

suggested by a few data points in two earlier studies, one

rural and one urban (7, 8). Cores from older urban lakes

sampled in this study (HAR, PAK, and NJOR) record the

presence of the 1950s peak and subsequent decreasing trend

but also document the reversal of the decreasing trend with

variable but increasing PAH concentration to the present

(1990s). This suggests that at older urban sites the improvements to sediment quality caused by changes in home heating

and power generation technology have since been overwhelmed by increases in other sources of PAHs.

The increasing trends in ∑PAH in all 10 water bodies over

the last 20-40 years have resulted in high to extremely high

concentrations in recently deposited sediments. To put the

concentrations of PAHs analyzed in this study into an

ecological perspective, the Canadian Sediment Quality

Guidelines for the Protection of Aquatic Life (27) for three

of the most prevalent PAHs are listed at the bottom of Table

3. Concentrations in surficial sediments (top of core) for all

10 lakes exceed the interim freshwater sediment quality

guidelines (ISQGssthe concentration below which adverse

effects are unlikely to occur) for these three compounds and

6 or more of the 10 lakes exceed the probable effect levels

(PELssthe concentration above which adverse biological

effects are expected to occur) for these compounds. For the

worst case, Newbridge Pond, concentrations of benzo[a]-

pyrene, fluoranthene, and pyrene are 14, 12, and 25 times

the PELs, respectively. Clearly, increases in PAHs in urban

water bodies have degraded sediment quality to the point

that it is an ecological concern.

The increase in PAHs in new urban settings cannot be

attributed solely to urbanization of the watersheds (Figure

3). This is most clearly illustrated by those watersheds in

which urban levels are stable (Table 1; BAL, HAR, NEW, and

TWN). The increase in PAH concentrations in these watersheds is, however, coincident with increases in automobile

use (Figure 4). Among the sources of PAHs related to

automobiles are tire wear, crankcase oil, roadway wear, and

car soot and exhaust (19, 30-32). Trends in ∑PAH from the

1970s to the present compare well to trends in VMT on

freeways and major arterial streets for the associated urban

areas for all sites and ages of urban development (Figure 4).

One intriguing possibility suggested by the results of this

study is that urban growth or sprawl outside the watershed

may adversely affect water quality within the watershed. This

is most evident when a watershed has undergone only a

relatively minor change in degree of urbanization. For

example, Austin, TX, is one of the most rapidly growing cities

in the country, but the majority of the growth there has

occurred around the fringes of the city and outside of the

watershed of Town Lake. This growth has contributed to

large increases in vehicle traffic in the Town Lake watershed;

traffic on MoPac Expressway, most of which lies within the

watershed and which crosses Town Lake, almost doubled

between 1990 and 1997 (33). The large increases in traffic

offer an explanation for why PAHs more than doubled in

Town Lake from 1975 to 1990 while percent urban land use

only increased by 5%. This suggests that urban sprawl in

outlying areas may affect traffic patterns and water quality

in the inner city.

The similar trends in VMT and PAHs in reservoir and lake

sediments seen here indicate that, in the absence of

reductions in vehicle-related PAH releases, ∑PAH will

continue to increase at rates comparable to those projected

for automobile use. Reduction of PAHs in the environment

will not be easily achieved. Although fluxes of several

contaminants (nitrogen oxides, airborne particulates <10

μm, carbon monoxide, and volatile organic compounds) in

transportation-related air emissions in the United States have

been decreasing since the 1970s (34), PAHs in the environment, as shown here, are continuing to increase coincident

with increasing VMT. This apparent contradiction may result

from the fact that there are several sources of vehicle-related

PAHs in addition to exhaust, including asphalt wear, tire

wear, and leaks and spills of engine oil. The multiple vehiclerelated sources, coupled with increasing dependence on the

automobile, highlight the complexity of reducing PAH

releases into the environment.

Acknowledgments

We thank Robert Eganhouse and Tom Lopes (U.S. Geological

Survey) for their many helpful comments. We also thank

Corey Stephens (U.S. Geological Survey) for compilation of

land-use data for the sampled watersheds. Mention of brand

names in the text is for identification purposes only and

does not constitute endorsement by the U.S. Geological

Survey.

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Received for review August 31, 1999. Revised manuscript

received July 6, 2000. Accepted July 18, 2000.

ES991007N

4070 9 ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 34, NO. 19, 2000

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