2000 Estuarine Research Federation 838

Estuaries Vol. 23, No. 6, p. 838–853 December 2000

Responses of Tidal Creek Macrobenthic Communities to the

Effects of Watershed Development

SCOTT B. LERBERG1,*, A. FREDERICK HOLLAND2, and DENISE M. SANGER2

1 Virginia Institute of Marine Science, School of Marine Science, P. O. Box 1346, Gloucester Point,

Virginia 23062 2 South Carolina Marine Resources Research Institute, 217 Fort Johnson Road, Charleston, South

Carolina 29422

ABSTRACT: This study examined the effects of watershed development on macrobenthic communities in tidal creeks

of Charleston Harbor, South Carolina, U.S. Two types of creeks were evaluated: upland creeks which drained watersheds

consisting of at least 15% terrestrial land cover, and salt marsh creeks which drained no upland habitat (i.e., only intertidal

habitat). Samples of macrobenthic organisms were taken along the longitudinal axis of twenty-three primary (first order)

tidal creeks. Water and sediment quality data were also collected including measurements of dissolved oxygen, salinity,

temperature, sediment characteristics, and toxic chemicals in the creek sediments. Hypoxic conditions occurred more

than 15% of the time in both reference and developed creeks and were a natural attribute of these systems. The most

severe and frequent hypoxia occurred in impacted salt marsh creeks. Salinity fluctuations were the greatest in developed

upland creeks and salinity range was identified as a potentially reliable indicator of the degree to which watershed

development has altered hydrodynamic processes. The creeks draining urban and industrial watersheds were degraded

environments characterized by watersheds with high (50%) levels of impervious surface, broad fluctuations in salinity,

severe hypoxia, and potentially toxic levels of chemicals in the sediment. These creeks had low macrobenthic diversity

and abundance and were numerically dominated by the oligochaete Monopylephorus rubroniveus in mud sediments, and

the polychaete Laeonereis culveri in sand sediments. Suburban watersheds had 15–35% impervious surface and creeks

draining them were exposed to frequent hypoxia and broad salinity fluctuations. The levels of chemical contaminants in

sediments of suburban and impacted salt marsh creeks were generally not different from the levels in reference creeks.

Macrobenthic diversity and abundance were higher for suburban and impacted salt marsh creeks than for urban and

industrial creeks. However, suburban and salt marsh impacted creeks were numerically dominated by a few pollution

indicative species including the oligochaetes M. rubroniveus and Tubificoides brownae and the polychaete L. culveri. These

creeks appear to be exhibiting early signs of degradation (e.g., a simplified food web). Two promising community-level

macrobenthic metrics for assessing environmental quality were identified: the proportional abundance of pollution indicative taxa, and the proportional abundance of pollution sensitive taxa. These indicators were significantly (p  0.05)

correlated with the salinity range, the level of chemical contaminants in sediments, and amount of impervious surface

in the watershed.

Introduction

Estuaries of the southeastern United States have

1.5–3.0 m tidal ranges, low relief, shallow depth,

and expansive intertidal habitats (Nummedal et al.

1977; Wiegert and Freeman 1990). Spartina alterniflora and Juncus roemerianus salt marshes and the

associated system of tidal creeks are characteristic

features of these environments (Teal 1962; Hackney et al. 1976; Wiegert and Freeman 1990). These

creeks provide nursery habitats for many species of

juvenile fish and crustaceans as well as feeding

grounds for wading birds and adult fish (Shenker

and Dean 1979; Weinstein et al. 1980; Wenner

1992; Wenner and Beatty 1993; Dodd and Murphy

1996). Hydrodynamic processes in tidal creeks also

trap fine-grained sediments, organic material, and

* Corresponding author: tele: 804/684-7371; fax: 804/684-

7399; e-mail: lerbergs@vims.edu.

chemical contaminants (Biggs et al. 1989) and

serve as both conduits and repositories for pollutants (Olsen et al. 1982; Horlick and Subrahmanyam 1983; Sanger et al. 1999a,b).

Over the next several decades, the watersheds

that drain tidal creeks in the southeastern United

States, as well as globally, are projected to experience a high rate of human development (Edwards

1989; Culliton et al. 1990; Cohen et al. 1997). The

construction of highway systems, residential housing, and industrial facilities that are projected to

accompany this development will alter the topography and drainage patterns. These alterations will

increase the rate and volume of freshwater inflow

as well as the amount of fine sediments, toxic

chemicals, organic materials, and nutrients introduced into estuaries (Driver and Troutman 1989;

Vernberg et al. 1992; Fulton et al. 1993; Holland

et al. 1996; Kucklick et al. 1997). Biological deg-
Watershed Development in Tidal Creeks 839

Fig. 1. Map of the Charleston Harbor Estuary showing the

locations of the tidal creeks sampled (mouth at approximately

3245N 7953W). Circles indicate reference creeks. Diamonds

indicate developed creeks.

radation associated with human-induced alterations to the ecosystem has been reported in many

freshwater river and stream environments (Karr

1991; Schueler 1994). If development proceeds in

an uncontrolled manner, the productivity and processes of tidal creek ecosystems may be irreversibly

degraded (Vernberg et al. 1992; Sanger et al.

1999a,b).

This study was part of a larger project (Tidal

Creek Project) evaluating the linkages between tidal creek environmental quality and human uses of

coastal watersheds (Holland et al. 1996). The Tidal

Creek Project used a comparative watershed approach that contrasted the physical, chemical, and

ecological characteristics of tidal creeks draining

relatively undeveloped watersheds with creeks

draining urban, suburban, and industrial watersheds. Study elements included water quality, sediment quality, macrobenthic communities, and

nekton communities. The specific objectives of this

study were to evaluate the impacts of watershed

development on macrobenthic communities and

evaluate linkages with environmental parameters

likely to affect them.

Materials and Methods

STUDY AREA AND SAMPLING DESIGN

Twenty-three tidal creeks located in Charleston

Harbor were sampled from July 5 to September 10,

1994 (Fig. 1). Eleven creeks were dominated by

urban, suburban, and/or industrial land cover (developed creeks), and eight drained undeveloped

forested watersheds (reference creeks). Four

creeks that drained predominately salt marsh vegetation were also sampled. Creeks were classified

into two subpopulations for analysis: upland and

salt marsh. Upland creeks were further classified

into four land use categories: forested or reference

(i.e.,  15% of the watershed as urban/suburban

land cover and some freshwater inflow); suburban

(45% urban/suburban land cover with a human

population density 5 but  20 individuals per

hectare); urban (i.e., 70% urban/suburban land

cover with a human population density of 20

individuals per hectare); and industrial (i.e., 

45% urban/suburban land cover and presence of

industrial facilities). Salt marsh creeks were either

located in regions of the Charleston Harbor Estuary far removed from any source of pollution (reference) or were located adjacent to highly developed (impacted) watersheds.

All creeks had a water depth of  10 cm at low

tide, and were 2–3 m deep at high tide. The upper

creek boundary was the point where depth at

mean high tide was about 1 m, or where an impassable obstacle (a dike or security fence) was

reached. The lower boundary was the point where

the creek converged with another water body, or

the depth in the channel exceeded 3 m at high

tide. The creeks represented the range of salinity

distributions, sediment characteristics, creek

lengths (231 to 1491 m), watershed sizes, land cover types, and levels of human disturbance that occur in Charleston Harbor (Table 1).

Creeks were stratified into 300-m reaches and

macrobenthic community samples were collected

randomly within each reach at approximately 1 m

below the mean high tide line (mid-tide level). Six

samples were collected with a 45.6-cm2 coring device to a depth of 15 cm and sieved through a 500-

m screen. All organisms retained on the screen

were identified to the lowest possible taxonomic
840 S. B. Lerberg et al.

TABLE 1. Summary of land use and watershed condition indicators for upland and salt marsh creeks.

Watershed Type Creek Name, Code

Linear Measurements

Watershed

Size (ha)

Creek

Length

(m)

Land Use Category

Forest

(ha)

Salt Marsh

(ha)

Water

(ha)

Barren/

Disturbed

(ha)

Agriculture

(ha)

Urban/

Suburban

(ha)

Watershed Condition

Indicators

Population

Density

(# ha1)

Impervious

Surface

(%)

Upland Creeks

Forested

Forested

Forested

Forested

Forested

Forested

Forested

Forested

Industrial

Beresford Creek, BF

Deep Creek, DP

Foster Creek, FT

Horlbeck Creek, HB

Lachicotte Creek, LH

Lighthouse Creek, LC

Long Creek, LI

Rathall Creek, RT

Diesel Creek, DL

25.1

64.7

16.0

237.5

13.1

37.3

412.4

72.1

105.3

535

586

585

895

436

899

1,446

1,200

784

5.2

40.0

3.4

184.6

1.9

2.7

243.1

30.3

0.0

19.9

12.0

12.6

9.1

11.2

31.6

17.7

35.3

32.8

0.0

0.0

0.0

3.6

0.0

0.0

0.0

0.0

0.0

0.0

0.0

0.0

17.9

0.0

3.0

43.8

6.5

8.9

0.0

12.7

0.0

22.3

0.0

0.0

101.8

0.0

0.0

0.0

0.0

0.0

0.0

0.0

0.0

6.0

0.0

63.6

0.0

0.0

0.0

0.1

0.0

0.0

0.3

0.6

6.9

0.0

4.1

0.0

1.8

0.0

0.0

2.2

0.8

50.0

Industrial

Industrial

Urban

Urban

Suburban

Suburban

Suburban

Suburban

Suburban

Suburban

Koppers Creek, KP

Shipyard Creek, SH

New Market Creek, NM

Vardell Creek, VR

Bull Creek, BL

Cross Creek, CC

Metcalf’s Creek, MC

Parrot Creek, PC

Shem Creek, SM

Yacht Club Creek, YC

116.6

279.1

187.2

68.7

442.8

312.9

129.7

147.2

427.8

69.1

880

826

1,167

450

1,395

536

874

1,491

1,174

593

0.0

14.1

3.0

0.0

83.5

28.6

17.4

23.6

61.5

4.6

11.6

10.2

26.6

9.5

20.5

5.2

7.0

14.5

9.3

8.2

0.0

0.0

0.0

0.0

15.6

2.4

0.0

1.4

4.2

1.0

12.3

74.5

17.9

0.0

3.5

41.0

17.1

30.6

66.8

1.8

0.0

0.0

0.0

0.0

0.0

8.9

0.0

7.8

0.0

0.0

92.7

180.4

139.7

90.5

319.7

226.8

88.2

69.3

286.0

53.5

3.3

8.2

21.1

31.8

12.7

10.6

8.9

7.0

15.7

7.1

52.7

49.4

51.2

70.4

28.9

26.4

30.8

19.5

34.5

14.9

Salt Marsh Creeks

Reference

Reference

Impacted

Impacted

Battery Simpkin Creek, BS

Grice Cove Creek, GC

Dill Creek, PI

Orange Grove Creek, OG

23.0

20.6

20.2

20.6

1,110

577

536

285

0.0

0.1

0.0

0.1

23.0

20.5

20.0

20.5

0.0

0.0

0.0

0.0

0.0

0.0

0.0

0.0

0.0

0.0

0.0

0.0

0.0

0.0

0.2

0.0

0.0

0.0

0.0

0.0

0.0

0.0

0.0

0.0

level and counted. Six surface (upper 2 cm) sediment samples were also collected and processed to

determine percent moisture, percent silts and

clays, and percent sand using the standard pipette

protocol modified from Plumb (1981).

A Hydrolab DataSonde 3 (DS3) water quality

monitoring system was deployed in the lower reach

of each creek for 2 to 4 d prior to sampling. The

DS3 was positioned approximately 5–10 cm above

the creek bottom and measured salinity, water temperature, dissolved oxygen concentration, pH, and

water depth at 30-min intervals.

Surface sediments (upper 2 cm) were also collected during the summer of 1995 for chemical

analyses at one randomly selected site from the upper-most and lower-most reach of each creek sampled in 1994. These samples were homogenized

then divided into three aliquots. One aliquot was

placed in an acid washed plastic jar for trace metal

analyses. The second aliquot was placed in a precleaned glass jar for organic contaminant analyses.

The third aliquot was stored in a plastic bag for

organic carbon analyses. The samples were then

immediately placed on ice and stored at 60C.

Analyses were performed for 14 trace metals, 24

polyaromatic hydrocarbons (PAHs), and 20 polychlorinated biphenyls (PCBs) by the National

Ocean Service Laboratory (National Oceanic and

Atmospheric Administration [NOAA]-Charleston,

South Carolina) using methods described by Sanger et al. (1999a,b). The contaminants measured

during the study include the majority of the contaminants routinely monitored by NOAA’s National Status and Trends Program (1991).

The boundary of each watershed was defined using topographic maps and Geographical Information System (GIS). The area was classified into land

use categories based on a modification of the Anderson Land Use Classification System (Anderson

et al. 1976). Human population statistics for each

watershed were obtained from the BerkeleyCharleston-Dorchester Council of Governments

(TIGER/Line Files 1992). The percent of each watershed that was impervious surface was determined by point sampling a minimum of 200 randomly selected points within each watershed on 1:

4,800 blackline aerial photographs (photographed

in 1989) and counting the number of points that

fell on roads, parking lots, roofs, or other impervious surfaces.

STATISTICAL ANALYSES

Abundance data for the numerically dominant

taxa was evaluated as the estimated number of individuals m2 (absolute abundance) or as the percent of the total infauna (relative abundance) col-
Watershed Development in Tidal Creeks 841

lected. Only the relative abundance data are presented in this paper. The absolute abundance data

can be found in Lerberg (1997).

A two-way analysis of covariance (ANCOVA) using SAS (1989) with all significant interaction

terms was used to contrast macrobenthic community characteristics among land use classes and between the upper-most and lower-most reaches of

creeks. Community measures used included the

number of species per sample, Shannon-Weiner

Diversity Index (H’), total number of individuals

per sample, percent dominance by most abundant

species, and the relative abundance of the numerically dominant oligochaetes (Monopylephorus rubroniveus, Tubificoides brownae, and Tubificoides heterochaetus), polychaetes (Laeonereis culveri and Streblospio benedicti), and nemerteans. Covariates evaluated

included the mean creek salinity, the silt-clay content of the sediments, and the percent of the time

the dissolved oxygen concentration was hypoxic (

28% saturation). Covariates were included in the

model if they were significant at p  0.10. Abundance data were transformed using a log10 transformation. Proportional data were transformed using an arcsine square root transformation. Estimates of least square means and pair-wise comparisons were obtained for each ANCOVA model. The

assumption that there were no significant differences in the slopes of the treatment lines was tested by determining if the covariate by main treatment effects interaction term was significant. The

equal slopes model is presented. All statistical analyses were considered significant at p  0.10.

This probability level (p  0.10) was selected

over a more standard p  0.05 to reduce the probability of making a Type II error. We compared our

results using both a 0.05 and 0.10 probability level

and found relatively few differences. Use of the p

 0.10 therefore provided for an increased likelihood of identifying potential ecologically important problems. In all cases we have provided the

observed p values and as much detail on the statistical testing as possible.

Wilcoxon rank sum and independent sample ttests were also used to compare the physical-chemical data between reference and developed creeks.

Associations (strength and direction) between environmental variables and macrobenthic community characteristics were evaluated using a Pearson’s product moment correlation coefficient (SAS

1989).

A mean effects range-low quotient (ERLQ) was

developed to provide a cumulative measure of

trace metal, PAH, and PCB contamination in the

sediments of these tidal creeks. This calculation involved dividing each concentration by the ERL value published in Long et al. (1995), summing the

ratio values, and dividing by the number of individual compounds included in the calculation (n

 23) (Long et al. 1998). The specific procedures

used and analytes included in the calculations are

described in Sanger (1998).

Results

LAND USE AND HUMAN POPULATION DATA

Watershed size in the population of creeks sampled ranged from 13.1 to 442.8 hectares (ha). The

drainage areas of developed upland creeks (mean

 203.4 ha) were generally larger than the drainage areas of the reference upland creeks (mean 

109.8 ha). The larger drainage basins had the most

complex land cover patterns. The four salt marsh

creeks had the smallest drainage areas (range 

20.2 to 23.0 ha, mean area  21.1 ha), and consisted of primarily one land cover type—estuarine

salt marsh. Population density in developed upland

watersheds ranged from 0 to 31.8 individuals ha1

(Table 1). The human population density in forested watersheds never exceeded 0.6 individuals

ha1. Creeks located in forested watersheds had little or no impervious surface in their drainage basins. Creeks in suburban and urban/industrial watersheds had 15–79% of their drainage area as impervious surface (Table 1). The amount of impervious surface in upland watersheds was significantly

correlated (p  0.0001, r2  0.62) with human

population density. When industrial watersheds

were excluded from this analysis the r2 value increased to 0.98.

PHYSICAL AND CHEMICAL DATA

Salinity varied within creeks, among creeks, and

among watershed classes (Fig. 2). Reference and

developed creeks encompassed a broad range of

salinities including euhaline (25–32 psu), polyhaline (18–25 psu), and mesohaline (5–18 psu) environments. Upstream-downstream salinity gradients (range  0–11.7 psu, mean  1.75 psu) were

small because most of the volume of creeks (ranging in length from 231 to 1,491 m) was displaced

on each tidal cycle. The mean salinity of upland

creeks was about five psu lower during low tide

than the comparable high tide salinity. Rainfall occurred shortly before or during almost every sampling event, with a measurable rain event occurring on 36 d of the 68-d sampling period. The average rainfall event resulted in 1.79 cm of rain.

During rain events (e.g., thunderstorms) large,

rapid declines in the low tide creek salinity occurred, even for the salt marsh creeks. Salinity fluctuations, as represented by the salinity range, were

significantly greater (p  0.01) in the suburban

and urban/industrial creeks than in the reference

(forested) upland creeks (Fig. 2). The salinity
842 S. B. Lerberg et al.

Fig. 2. The mean, median, and salinity range for each creek collected two to four days prior to each sampling event. The top of

the bar represents the mean value and the shaded square represents the median value. The vertical line represents the range. The

mean salinity range (mean SRw 	 SE) for each watershed class is as follows: Forested (6.9 psu 	 1.2), Industrial (12.8 psu 	 2.9),

Urban (14.5 psu 	 1.4), Suburban (14.0 psu 	 3.2), Reference Salt Marsh (14.6 psu 	 5.5), and Impacted Salt Marsh (6.9 psu 	

1.9).

range of upland creeks was positively correlated (p

 0.05) with the size of the drainage basin.

Dissolved oxygen (DO) values varied with time

of the day and tidal stage. In general, highest DO

levels (% saturation) occurred in mid-afternoon.

The lowest DO levels occurred during early morning low tides (Fig. 3). Mean DO levels were similar

among watershed classes (Fig. 4). Values representing supersaturated conditions occurred during

midday in upland and salt marsh creeks. Dissolved

oxygen levels in both reference and developed

creeks were frequently below 28% saturation (hypoxic). The percent of time upland creeks exhibited

low DO levels was similar among watershed classes

(Fig. 5, p 0.10).

Mean creek water temperatures were 25C.

Extreme values occurred during daytime low tides

(32C). No significant differences in water temperature occurred among creeks. The pH of creek

water tended to be lower in forested (upland reference) creeks (7.1 versus 7.3) probably because

of the higher levels of humic acids which are associated with runoff from forested environments.

Sediments varied in a heterogeneous manner

along the lengths of creeks. Few consistent upstream-downstream gradients in the silt-clay content were found. Exceptions included the downstream increase in silt-clay content in Shem, Metcalf, and Diesel creeks. Silt-clay content in Parrot

and Shipyard creeks exhibited the reverse pattern.

Greater amounts of fine sediments were found in

impacted salt marsh creeks than reference salt

marsh creeks. Forested creeks were characterized

by more fine sediment than suburban, industrial,

or urban creeks. The urban creek class contained

the greatest amount of sand sediments.

Surface sediments in creeks draining industrial

and urban watersheds contained significantly (p 

0.0001) greater amounts of chemical contaminants

than other creek classes particularly in their uppermost reaches (Fig. 6). A complete summary and

analysis of the contaminant data is presented in

Sanger et al. (1999a,b).

MACROBENTHIC COMMUNITY DATA

A total of 8,034 individuals representing 97 taxa

were obtained from the 417 samples. The assemblage was dominated by annelid worms, particularly oligochaetes (mean  2,208 individuals m2)

and polychaetes (1,811 individuals m2). These

taxa comprised 95.9% of the fauna. Molluscs (71

individuals m2) and crustaceans (45 individuals

m2) accounted for only 2.8% of the fauna (Tables

2 and 3).

Forty-four of the 97 taxa collected occurred in

only one or two samples. Nine taxa (M. rubroniveus,

S. benedicti, Heteromastus filiformis, L. culveri, T. heterochaetus, T. brownae, Capitella capitata, Tharyx cf.

acutus, and Neanthes succinea) accounted for 90%

of the specimens collected (Table 3).
Watershed Development in Tidal Creeks 843

Fig. 3. Summary of dissolved oxygen and depth data for a

representative (a) forested creek (Long Creek sampled July 24–

July 28, 1994), (b) suburban creek (Shem Creek sampled July

18–July 21, 1994), (c) industrial creek (Diesel Creek sampled

July 16–July 20, 1994), and (d) salt marsh impacted creek (Dill

Creek sampled July 22–July 25, 1994).

Fig. 5. The mean percent of time dissolved oxygen concentrations were less than 28% for creeks composing each watershed class. Error bars represent 	 SE.

Fig. 4. The mean, median, and range of dissolved oxygen data for each creek two to four days prior to each sampling event. The

top of the bar represents the mean value and the shaded square represents the median value. The vertical line represents the range.

MACROBENTHIC COMMUNITY RESPONSE TO NATURAL

ENVIRONMENTAL VARIATION

Few taxa exhibited abundance differences between the upper-most and lower-most creek reaches. The relative abundances of the oligochaete M.

rubroniveus (p  0.0079), and the polychaete C.

capitata (p  0.0403) were significantly higher in

the upper-most creek reaches than the lower-most

reaches. Conversely, the abundance of the polychaete H. filiformis was significantly higher (p 

0.0001) in the lower-most reaches. When upstream-downstream gradients in abundance were
844 S. B. Lerberg et al.

Fig. 6. Summary of the ERLQ (effects range-low quotient)

values for the upper and lower reach of each watershed type.

Error bars represent 	 SE.

TABLE 2. Summary of the macrobenthic data for upland and

salt marsh creeks by major taxonomic group (n  417 cores).

Major Taxa

No. of

Individuals

Collected

Average Density

(# m2)

% Total

Abundance

Oligochaeta

Polychaeta

Gastropoda

Nemertinea

Crustacea—Others*

Crustacea—Decapoda

Platyhelminthes

Pelecypoda

Hirudinea

Hemichordata

Hydrozoa

Sipuncula

4,198

3,443

132

130

61

24

20

19

4

1

1

1

2,208

1,811

69

68

32

13

11

10

2

1

1

1

52.25

42.85

1.64

1.62

0.76

0.30

0.25

0.24

0.05

0.01

0.01

0.01

* This group includes Amphipods, Isopods, Tanaids, Ostracods, and Mysids.

found, the actual amount of variation accounted

for by longitudinal distance was relatively small (

5% of the total variation).

The salinity covariate was significant (p  0.10)

for 9 of the 14 metrics evaluated (Table 4) and was

the most significant term in the ANCOVA model

for five of these analyses accounting for at least

10% of the model variance. Metrics that were positively associated with salinity included the number

of taxa, the total number of individuals, and the

relative abundance of all polychaetes, H. filiformis,

and S. benedicti. The relative abundance of all oligochaetes, T. heterochaetus, L. culveri, and the nemerteans decreased with increasing salinity.

The silt-clay covariate was significant for 7 of the

14 metrics evaluated (Table 4) and was the most

significant term in the ANCOVA model for three

metrics. However, the silt-clay covariate generally

accounted for  2% of the variance in the ANCOVA models. The relative abundance of the polychaete L. culveri was the only metric evaluated that

had substantial amounts of variance (14%) associated with the sediment covariate. Metrics that

were negatively correlated with the sediment siltclay content included the number of taxa per sample, the relative abundance of all polychaetes, and

the relative abundance of L. culveri, C. capitata, and

T. heterochaetus. The total number of individuals per

sample and the relative abundance of M. rubroniveus were positively correlated with silt-clay content.

The dissolved oxygen covariate was significant

for 6 of the 14 metrics evaluated (Table 4) and was

the most significant term in the ANCOVA model

for three metrics. However, it accounted for small

(2–5%) amounts of the variance. Metrics that decreased in value with decreasing dissolved oxygen

levels were the number of taxa per sample, and the

relative abundance of the polychaetes, S. benedicti

and L. culveri. Metrics that increased in value with

increasing exposure to low dissolved oxygen were

the total number of individuals, the relative abundance of all oligochaetes, and the relative abundance of M. rubroniveus.

MACROBENTHIC COMMUNITY RESPONSES TO

DEVELOPMENT

Tests comparing macrobenthic community metrics among watershed-types and reaches, and the

identification of interactions between watershed

and reach effects are presented in Table 5. Results

for the upland creek classes are discussed first followed by a discussion of salt marsh creeks. Data for

industrial and urban creek classes were combined

for these analyses because of the similarity in the

amounts of impervious surface for the two watershed classes. This combination also increases the

power of significance tests to detect differences between this new urban/industrial watershed class

and other creek classes.

The number of taxa per sample was significantly

higher in creeks located in the forested (p 

0.0566) and suburban (p  0.0199) watershed classes than in creeks draining urban/industrial watersheds. This pattern resulted because crustaceans

and bivalves were reduced in creeks draining urban and industrial watersheds. Rare taxa that were

not found in the industrial and urban creeks included T. cf. acutus, Fabricia sp., Potamilla sp., Leptochelia sp., Melita nitida, Gammarus mucronatus, Cyathura burbancki, Macoma tenta, Tellinidae, and Polymesoda caroliniana.

The total number of individuals per sample was

higher in upland creeks draining suburban watersheds than in creeks draining watersheds dominated by forested (p  0.0032) or urban/industrial (p
Watershed Development in Tidal Creeks 845

TABLE 3. Macrobenthic data collected from the upper-most and lower-most reaches of upland creeks and salt marsh creeks. Key to

Major Taxonomic Groups: A  Amphipoda, G  Gastropoda, HE  Hemichordata, I  Isopoda, M  Mysidacea, N  Nemertinea,

O  Oligochaeta, OS  Ostracoda, P  Polychaeta, PE  Pelycepoda, PLAT  Platyhelminthes, S  Sipuncula, T  Tanaidacea. Key

to Species Code: PI  Pollution Indicative, PS  Pollution Sensitive, U  Ubiquitous, R  Rare.

Scientific Name

Species

Code

Forested (n  93)

Avg. No.

m2 % Total

Suburban (n  71)

Avg. No.

m2 % Total

Urban/Industrial

(n  56)

Avg. No.

m2 % Total

Marsh Ref. (n  24)

Avg. No.

m2 % Total

Marsh Imp. (n  17)

Avg. No.

m2 % Total

Monopylephorus rubroniveus (O) PI 787.6 22.6 1,868.7 44.2 806.7 36.1 1,928.0 28.6 7,198.2 84.8

Tubificoides heterochaetus (O) PS 573.0 16.4 262.5 6.2 50.9 2.3

Heteromastus filiformis (P) PS 474.0 13.6 139.0 3.3 82.2 3.7 849.8 12.6 270.9 3.2

Streblospio benedicti (P) PS 466.9 13.4 444.8 10.5 246.7 11.0 2,412.3 35.7 528.9 6.2

Capitella capitata (P) U 285.3 8.2 40.2 0.9 231.0 10.3 18.3 0.3

Neanthes succinea (P) U 207.5 6.0 139.0 3.3 78.3 3.5 173.6 2.6 38.7 0.5

Laeonereis culveri (P) PI 113.2 3.2 417.0 9.9 513.0 22.9

Nemertinea (N) PS 108.5 3.1 40.2 0.9 70.5 3.2 64.0 0.9 64.5 0.8

Tubificidae (O) U 84.9 2.4 83.4 2.0 39.2 1.8 100.5 1.5 90.3 1.1

Tubificoides brownae (O) PI 82.5 2.4 525.1 12.4 82.2 3.7 721.9 10.7 270.9 3.2

Tharyx cf. Acutus (P) PS 63.7 1.8 30.9 0.7 246.7 3.7

Fabricia sp. A (P) R 49.5 1.4

Potamilla sp. (P) R 42.4 1.2 91.4 1.4

Tubificoides wasselli (O) R 25.9 0.7 3.1 0.1

Leitoscoloplos fragilis (P) R 21.2 0.6 27.8 0.7 3.9 0.2 27.4 0.4 12.9 0.2

Monopylephorus irroratus (O) R 16.5 0.5 37.1 0.9 9.1 0.1

Polydora cornuta (P) R 14.1 0.4 18.5 0.4 27.4 0.4

Leptochelia sp. (T) R 11.8 0.3

Ostracoda sp. (OS) R 11.8 0.3 6.2 0.1

Eteone heteropoda (P) R 7.1 0.2 9.1 0.1

Mediomastus californiensis (P) R 7.1 0.2 3.9 0.2

Cassidinidea lunifrons (I) R 4.7 0.1 18.5 0.4 3.9 0.2

Melita nitida (A) R 4.7 0.1 9.3 0.2

Amphitrite ornata (P) R 2.4 0.1

Corbicula contracta (PE) R 2.4 0.1

Cyathura burbancki (I) R 2.4 0.1 6.2 0.1 9.1 0.1

Glycera americana (P) R 2.4 0.1 3.1 0.1

Hesionidae (P) R 2.4 0.1 3.1 0.1 3.9 0.2

Macoma tenta (PE) R 2.4 0.1 6.2 0.1

Sipuncula (S) R 2.4 0.1

Tellinidae (PE) R 2.4 0.1

Websterinereis sp. (P) R 2.4 0.1

Turbellaria (PLAT) R 18.5 0.4 3.9 0.2

Bocardiella sp. A. (P) R 15.4 0.4

Mysidopsis bigelowi (M) R 12.4 0.3

Gastropoda (G) R 9.3 0.2

Monticellina dorsobranchialis (P) R 9.3 0.2 18.3 0.3

Paraprionospio pinnata (P) R 9.3 0.2

Polydora socialis (P) R 6.2 0.1

Gammarus mucronatus (A) R 3.1 0.1

Hemichordata (HE) R 3.1 0.1

Lumbrineridae (P) R 3.1 0.1

Mysidopsis furca (M) R 3.1 0.1

Polymesoda caroliniana (PE) R 3.1 0.1

Prionospio sp. (P) R 3.1 0.1

Sphaeorsyllis longicauda (P) R 3.1 0.1 9.1 0.1

Corbiculidae (PE) R 3.9 0.2

Sabellidae (P) R 3.9 0.2 9.1 0.1

Terebellidae (P) R 3.9 0.2

Corophium lacustre (A) R 9.1 0.1

Gemma gemma (PE) R 9.1 0.1

Mulinia lateralis (PE) R 9.1 0.1

Acanthohaustorius intermedius (A) R 12.9 0.2

 0.0002) land cover (Table 5). Most of the higher

abundances found in suburban creeks were associated with the oligochaete fauna. The percent of

the macrobenthic assemblage represented by oligochaetes and the relative abundance of the two

most abundant oligochaete species, M. rubroniveus

and T. brownae, were significantly higher in the suburban creek class than in the forested or urban/
846 S. B. Lerberg et al.

TABLE 4. Summary of the influence of environmental factors used as covariates in the ANCOVA models on community characteristics. A plus indicates a positive relationship between the covariate and the respective community characteristic. A minus sign indicates

a negative relationship.

Habitat/Exposure Indicators

Salinity

(mean from DS3)

Sediment Type

(% silts and clays)

Dissolved Oxygen

(% time below 28% saturation)

Community Parameters

Species Richness (# sample1) 


Shannon-Weiner Diversity (H)

Total Number of Individuals (# sample1) 




Dominance by Most Abundant Species (% abundance

sample1)

Dominant Taxa

Polychaetes 
 

Heteromastus filiformis

Streblospio benedicti





 

Laeonereis culveri

Capitella capitata

 





Oligochaetes  


Monopylephorus rubroniveus

Tubificoides brownae

Tubificoides heterochaetus 









Nemertinea 

TABLE 5. Summary of ANCOVA results evaluating the effects of watershed type and longitudinal position (reach) on community

parameters. Values shown are p levels. Pairwise comparisons were analyzed using least square means. Treatments not connected by a

common underline differed at p  0.10. SR  salt marsh reference, SI  salt marsh impacted, U  urban/industrial, SU  suburban,

and F  forested (upland reference), ns  not significant. An asterisk indicates that the models for these species were accomplished

using data for only creeks located on upland watersheds.

Macrobenthic Community Parameters

Significant

Covariates

Watershed

Type Reach Interaction Pairwise Comparisons

Number of taxa (# sample1) DO, sal, mud 0.0084 ns ns SR SU F SI U SR SU F SI U

Total number of individuals (# sample1) 0.0001 ns 0.0439 SI SR SU F U

Shannon-Weiner Diversity (H) DO, sal, mud 0.0082 ns ns SR F SU U SI SR F SU U SI

Dominance by most abundant species (% abundance sample1) 0.0145 ns 0.0433 SI SU U SR F SI SU U SR F

Oligochaetes (% abundance sample1) DO, sal 0.0002 ns ns SI SU SR U F SI SU SR U F

Monopylephorus rubroniveus mud, DO 0.0001 0.0079 ns SI SU SR U F SI SU SR U F

Tubificoides brownae 0.0001 ns ns SR SU SI U F

Tubificoides heterochaetus* sal, mud 0.0001 ns ns F SU U

Polychaetes (% abundance sample1) sal, mud ns ns ns SR S U F U SI U F U SI

Streblospio benedicti sal, DO 0.0609 ns ns SR F SI U SU SR F SI U SU

Laeonereis culveri* mud, DO, sal ns ns 0.0004 U SU F 

Nemertinea (% abun sample1) sal 0.0047 ns ns F SR U SI SU F SR U SI SU

industrial classes (Table 5). Abundances of M. rubroniveus and T. brownae were lowest in creeks draining forested watersheds.

The relative abundance of the polychaete L. culveri was significantly higher (p  0.0592) in creeks

draining urban/industrial watersheds than in

creeks draining forested watersheds. The abundance of this species in creeks draining suburban

watersheds was intermediate and not significantly

different (p 0.10) from that in creeks draining

urban/industrial or forested watersheds.

Several other taxa had abundance patterns that

appeared to be negatively associated with increasing degrees of watershed development. For example, the relative abundances of T. heterochaetus (p

 0.0001), S. benedicti (p  0.0416), and the nemerteans (p  0.0271) were significantly higher in

the forested creek class than in the suburban and

urban/industrial creek classes (Table 5).

Engle and Summers (1998) and Van Dolah et al.

(1999) suggest that the H measure of species diversity is positively correlated with estuarine envi-
Watershed Development in Tidal Creeks 847

TABLE 6. Summary of the responses of the macrobenthic community to watershed development. Values presented for mean number

of taxa per sample, Shannon-Weiner diversity (H), % dominance by most abundant species, % relative abundance of rare taxa, %

relative abundance of pollution indicative species, % relative abundance of potentially pollution sensitive taxa, % relative abundance

of ubiquitous taxa, and mean total number of individuals per sample.

Macrobenthic Community Measures

Upland Creeks

Forested Suburban Urban/Industrial

Salt Marsh Creeks

Reference Impacted

Number of taxa per sample 3.4 3.6 2.5 4.9 2.8

Shannon-Weiner Diversity (log10) 0.37 0.36 0.30 0.48 0.21

% dominance 48 58 52 48 80

% abundance of rare species 5.8 4.6 0.7 1.6 0.1

% pollution indicative species 27.3 64.8 60.5 38.7 84.3

Monopylephorus rubroniveus 21.9 43.1 34.8 28.2 81.2

Laeonereis culveri 3.1 9.6 22.1 0.0 0.0

Tubificoides brownae 2.3 12.1 3.5 10.5 3.1

% pollution-sensitive species 46.8 21.2 19.4 52.2 9.8

Tubificoides heterochaetus 15.9 6.1 2.2 0.0 0.0

Streblospio benedicti 13.0 10.3 10.6 35.2 6.0

Nemertinea 3.0 0.9 3.0 0.9 0.7

Heteromastus filiformis 13.2 3.2 3.5 12.4 3.1

Tharyx cf. acutus 1.8 0.7 0.0 3.6 0.0

% ubiquitous taxa 13.7 4.1 13.3 2.8 0.4

Capitella capitata 7.9 0.9 10.0 0.3 0.0

Neanthes succinea 5.8 3.2 3.4 2.5 0.4

Total number of individuals 3,348 4,294 2,562 6,402 8,423

ronmental quality at the regional scale. This study

found that H was not significantly different among

upland creek classes (Table 5). Diversity (H) values did tend to decrease as the degree of development increased. The dominance measure evaluated (relative abundance of the most abundant

species) was significantly higher (p  0.0381) in

suburban creeks than in forested creeks. Dominance values for urban/industrial creeks were intermediate between the values for suburban and

forested creeks, but were not significantly different

(p 0.10) from either class of creek (Table 5).

The salt marsh reference creeks had significantly

higher numbers of taxa (p  0.0924) and diversity

(p  0.0009) than the impacted salt marsh creeks.

Much of this difference was due to the absence of

rare taxa, particularly crustaceans and bivalves,

from impacted salt marsh creeks. Rare taxa that

were not found in impacted salt marsh creeks included T. cf. acutus, Eteone heteropoda, C. burbancki,

Corophium lacustre, Mulinia lateralis, and Gemma gemma. In addition, the impacted salt marsh creeks

had significantly higher dominance values than the

other watershed classes sampled (p  0.0555).

Macrofaunal abundance was consistently higher

for salt marsh creeks than for upland creeks, although there was no difference in total macrofaunal abundance between the two salt marsh creek

populations.

The oligochaete M. rubroniveus was the numerically dominant macrobenthic organism in salt

marsh creeks. Densities of this species were significantly higher (p  0.0001) in impacted salt marsh

creeks than in reference salt marsh creeks. The oligochaete, T. brownae, exhibited the opposite pattern and had significantly (p  0.0556) higher

abundances in reference salt marsh creeks than in

impacted salt marsh creeks. Two species, L. culveri

and T. heterochaetus, that were numerically dominant in upland creeks were not found in salt marsh

creeks. Both of these species appear to prefer sand

habitats and the high levels of fine sediments that

occur in the salt marsh creeks may have effectively

excluded them from these environments.

A summary of the changes in relative abundance

of numerically dominant species and communitylevel metrics among watershed classes is presented

in Table 6. The relative abundance of three of the

numerically dominant species (M. rubroniveus, T.

brownae, and L. culveri) increased with increasing

levels of watershed development and pollution

stress. The two oligochaete species, M. rubroniveus

and T. brownae, attained their highest abundance

in the suburban creeks (Table 6). The high abundance of L. culveri in urban/industrial creeks may

have resulted from the higher proportion of sand

sediments in this class of creeks (e.g., Vardell and

Shipyard).

The oligochaete T. heterochaetus, the nemerteans,

and the polychaete S. benedicti declined in abundance with increasing levels of watershed development (Tables 5 and 6). Most research on ma-
848 S. B. Lerberg et al.

Fig. 7. Relationships between (a) human population density,

(b) salinity range, and (c) total chemical exposure and the

amount of impervious surface on the watershed. The coefficient

of determination (r2) and p-value are presented.

Fig. 8. Relationships between two indices of macrobenthic

community measures and the amount of impervious surface in

the watershed. Only data from the upland creeks were used.

Potentially pollution sensitive taxa (a) include Tubificoides heterochaetus, Streblospio benedicti, Heteromastus filiformis, Tharyx cf. acutus, and Nemertinea. Pollution indicative taxa (b) include Monopylephorus rubroniveus, Tubificoides brownae, and Laeonereis culveri.

The coefficient of determination (r2) and p-value are presented.

crobenthic indicators suggest that S. benedicti is relatively tolerant to pollution and list this polychaete

as a pollution indicative organism (Pearson and

Rosenberg 1978; Levin 1984; Weisberg et al. 1990;

Engle et al. 1994). However, our findings and the

experiments of Sarda et al. (1996) suggest that S.

benedicti is not as tolerant to chronic pollution exposure or hypoxia as the oligochaete M. rubroniveus.

Other species that are potentially useful as indicators of pollution stress include T. cf. acutus and

H. filiformis. Tharyx cf. acutus decreased in abundance as the degree of watershed development increased, however, this species only occurred in

high polyhaline and euhaline environments. Heteromastus filiformis had lower relative abundance in

creeks located on developed watersheds (Table 6).

These differences were not significant (p 0.10)

because of high within-habitat variability in creeks

with forested watersheds. Lower abundances of

these species in developed creeks do not unambiguously indicate the degree of development stress.

In contrast, N. succinea and C. capitata were ubiquitous organisms which showed little differences in

abundance with the type and degree of watershed

development and appear to be not useful as indicators of pollution stress (Table 6).

The amount of impervious surface in a watershed has been found to be a good indicator of

watershed condition by a number of studies

(Schueler 1994; Arnold and Gibbons 1996). In this

study there was a significant relationship between

the amount of impervious surface in the watershed

sampled and the human population density, the

salinity range as determined from the DS3 data,

and the cumulative level of chemical contaminants

in creek sediments as estimated by the ERLQ (Fig.

7). The slopes of all the lines were positive and

significantly different from zero (p  0.025).

Relationships between the amount of impervious surface in the watershed and two promising

macrobenthic community-level metrics, the percent abundance of pollution indicative species (L.

culveri, M. rubroniveus, and T. brownae) and the percent abundance of potential pollution sensitive

species (T. heterochaetus, S. benedicti, Nemertinea, T.

cf. acutus, and H. filiformis), are shown in Fig. 8.
Watershed Development in Tidal Creeks 849

The relative abundance of potentially pollution

sensitive taxa was negatively associated with the

amount of impervious surface. The slope of the

line was significantly different from zero (p 

0.0027) with a marked decline in abundance occurring when the percent impervious surface exceeded 30%. The relative abundance of pollution

indicative taxa increased with increasing amounts

of impervious surface. The slope of this line was

also significantly different from zero (p  0.0087).

Discussion

Tidal creeks are the primary linkage between upland and estuarine environments. Tidal creeks

function as repositories and conduits for sediments, organic materials, and chemical pollutants

(Olsen et al. 1982; Sanger et al. 1999a,b). The

highest levels of toxic chemicals were generally

found in the upper reaches of industrial and urban

creeks and were mainly associated with historical

industrial point source discharges. Sediment contaminant concentrations in suburban creeks were

not substantially different from values reported for

forested creeks and did not attain levels likely to

affect living organisms. Shem Creek was a notable

exception in that it had elevated levels of both sediment trace metal and organic contaminants. This

was not surprising since the Shem Creek watershed

has a relatively high human population density (

15 individuals ha1), receives runoff from commercial developments and a major highway corridor, and is used as a dockage facility for over 50

commercial fishing vessels and a dry stack boat

storage yard (Sanger 1998; Sanger et al. 1999a,b).

Suburban development produces few sources for

trace metal input and generally low levels of organic contaminant input to tidal creeks (Fortner

et al. 1996; Sanger et al. 1999a,b).

Sediments in salt marsh creeks were predominately fine grained pluff muds. Creek bottom habitats in upland creeks, particularly developed upland creeks, were frequently composed of large

amounts of sand. Soils in the upland portion of

the watersheds studied were predominately sandy.

In addition, fine sediments may have been scoured

from the creek channels due to altered hydrography and deposited on the marsh surface and/or

in the deeper portions of the tidal creek and estuary.

Tidal creek DO dynamics were complex. Concentrations in both upland (forested and developed) and salt marsh (reference and impacted)

creeks were frequently hypoxic ( 28% saturation)

indicating exposure to low DO is a natural phenomenon in these systems. The observed DO levels were frequently within ranges reported to adversely affect macrobenthic distributions (Gaston

1985; Diaz and Rosenberg 1995). Impacted salt

marsh creeks had the lowest and most variable DO

levels. Dissolved oxygen dynamics in these creeks

were similar to that reported for poorly flushed

dead-end canals (Baca et al. 1988; Maxted et al.

1997). Developed creeks generally experienced

more frequent exposure to low DO concentrations

and had more pronounced tidal periodicity in DO

than reference creeks. For a more detailed description of the DO dynamics of the tidal creeks sampled refer to Wenner et al. (1999).

The tidal creeks sampled tended to have lower

biodiversity relative to comparable estuarine habitats (Engle et al. 1994; Weisberg et al. 1997). In

both upland and salt marsh creeks, species richness, number of taxa per sample, and percentage

of rare species decreased as watershed development and environmental impact increased. The

creeks with the fewest species, particularly rare species, were also characterized by the most severe

and prolonged exposure to the low dissolved oxygen levels, the highest levels of sediment contamination, and the largest salinity fluctuations. This

finding is supported by a number of other studies

(Pearson and Rosenberg 1978; Broom et al. 1991;

Engle et al. 1994).

Tidal creeks were numerically dominated by a

few species of stress tolerant oligochaetes and polychaetes. Many of these species have life history

characteristics which allow them to maintain high

abundances in organically enriched environments

that experience periodic hypoxia including rapid

growth and development, multiple modes of reproduction including asexual reproduction, and

the ability to recruit throughout much of the year

(Grassle and Grassle 1974; McCall 1977; Pearson

and Rosenberg 1978; Giere and Pfannkuche 1982;

Levin 1984; Levin and Huggett 1990; LaSalle et al.

1991). Sarda et al. (1996) present evidence that

long-term chronic exposure to organic enrichment

and associated environmental conditions provide a

competitive advantage to oligochaetes (e.g., Monopylephorus evertus), particularly during fall and winter when juvenile fish and crustacean predators are

not abundant. These authors speculate that food

resources and predation limit the abundances of

oligochaetes during summer and suggest opportunistic polychaetes (e.g., S. benedicti) may be better

adapted to respond to disturbance from spring

and summer predation. McCann and Levin (1989)

further indicate that mobile subsurface deposit

feeding oligochaetes (e.g., M. evertus) inhibit

growth of sedentary tubiculous polychaetes (e.g.,

S. benedicti) but not their recruitment. Additional

information on the seasonal distribution of tidal

creek macrobenthos and the effects of predation

on recruitment patterns is needed. For example,
850 S. B. Lerberg et al.

the data in Sarda et al. (1996) and Sanger (1998)

suggest that winter sampling may provide a more

sensitive measure of the effects of watershed development on macrobenthic communities.

The ecological role of oligochaetes in estuarine

systems is poorly understood (Diaz 1980; McCann

and Levin 1989), although oligochaetes have high

nutritional value and are a major food item in the

diet of juvenile fish and crustaceans (Hunter and

Arthur 1978; Van den Broek 1978; Giere and

Pfannkuche 1982; Wiltse et al. 1984; McCann and

Levin 1989). Many oligochaetes contain lipids

which are critical for shrimp reproduction (Middleditch et al. 1979). The dominant oligochaete in

South Carolina tidal creeks, M. rubroniveus, is ubiquitous in estuaries worldwide (Pfannkuche 1978)

but little information is known about its environmental tolerances. A closely related Pacific species,

Monopylephorus cuticulatus, is tolerant to anoxia,

trace metal contamination, organic enrichment,

and pulp mill effluents (Chapman et al. 1982).

Our data suggest the same is true for M. rubroniveus. Additional research is needed to test this

speculation.

Watersheds with urban and industrial development were associated with high levels of impervious surface (50%) and had severe impacts on

tidal creek habitats including alterations to the hydrodynamic processes (as indicated by salinity

range), exposure to levels of chemical contaminants that cause mortality, altered growth, and reduced reproduction, and exposure to more frequent and severe hypoxia. The macrobenthic community in urbanized and industrialized creeks, especially the upper reaches, was characterized by

low diversity, low numbers of rare and pollution

sensitive species, and low macrobenthic abundances (Table 6). The physical/chemical and the biological impacts reported by this study indicate a

severe impact to the nursery functions of tidal

creeks in urban/industrial watersheds.

Suburban creeks had higher densities of pollution tolerant macrobenthic organisms, particularly

the oligochaetes M. rubroniveus and T. brownae (Table 6). The levels of sediment contamination and

hypoxia in these creeks were, however, generally

not different than those in the forested upland

creeks. The dominance in suburban creeks by a

few species represents a trend towards a simplified

food web and may be an early warning of degradation (Diaz 1992). Simplification of food webs

may cause irreversible changes in ecosystem processes (Fretwell 1987; Pimm and Kitching 1987).

These findings suggest that macrobenthic responses in suburban creeks may be a more sensitive and

reliable indicator of adverse effects than water or

sediment quality data.

The major physical alteration found for impacted salt marsh creeks was hypoxia. Elevated levels

of sediment contaminants were not found in these

creeks. Impacted salt marsh creeks were characterized by a high abundance of M. rubroniveus, a low

abundance of S. benedicti and H. filiformis, and very

low species diversity.

Salinity fluctuations in tidal creeks appeared to

have potential as a measure of the degree to which

development had altered hydrodynamic processes

including the rate and volume of non-point source

freshwater runoff. This metric was significantly correlated (p  0.005) with both of the promising

community-level indicators identified. Changes in

stream flow regime, resulting from modifications

of the land surface with changing land use, is one

of the primary environmental factors affecting the

biological integrity of aquatic biota in freshwater

stream environments (Karr and Dudley 1981). Additional research in estuarine systems is needed to

evaluate the effects of storm events, watershed size,

and tidal creek type on this potential indicator of

development impacts. Research is also needed to

determine how much of the variation in macrobenthic response, especially in suburban creeks, is

due to alterations to salinity distributions. If the

salinity range indicator were proven to be reliable,

high quality salinity data could be collected routinely and economically with any number of available remote water quality monitoring systems.

In freshwater stream environments, numerous

studies (Miller and Mattraw 1982; Brown 1988;

Schueler 1994; Arnold and Gibbons 1996) suggest

the amount of impervious surface is a good measure of the degree to which human development

has altered watershed condition. In these studies,

when the amount of impervious surface exceeded

about 10%, the ecological condition of streams was

impacted. When the amount of impervious surface

exceeds about 30%, streams became permanently

degraded (Arnold and Gibbons 1996). Our findings suggest a similar pattern occurs in the tidal

creeks. Clearly definable impacts to tidal creek water and sediment quality and macrobenthic communities did not occur until the percent impervious surface exceeded 30%. This is not surprising

since South Carolina estuaries have relatively large

tidal ranges (1 m) and large tidal mixing.

The two promising community-level indicators

developed by this study based on the data collected

(i.e., the relative abundance of pollution indicative

taxa and the relative abundance of pollution sensitive taxa) were more sensitive indicators of ecological impact than other direct benthic metrics or

measures of sediment and water quality. Other

benthic metrics include the abundance of individual numerically dominant species and frequently
Watershed Development in Tidal Creeks 851

used community-level metrics of condition (e.g.,

H, McIntosh’s Dominance Index). Numerous

studies have found community-level indicators or

multi-metric indices (e.g., index of biotic integrity)

developed from broad scale comprehensive data to

be sensitive and reliable measures of ecological

condition (e.g., Pearson and Rosenberg 1978; Karr

1981; Bilyard 1987; Warwick and Clark 1991; Diaz

1992; Engle et al. 1994; Weisberg et al. 1997). Additional research is needed to validate these indicators at regional scales.

ACKNOWLEDGMENTS

We wish to thank C. Biernbaum, B. Roumillat, and C. Wenner

for providing input into the design and analysis of this work

and in the preparation of the manuscript. We also wish to recognize the efforts of George Riekerk, Lynn Zimmerman, Anita

Black, Elizabeth Daniel, Jackie Thompson, and Bonnie Whitlock

for participating in fieldwork, sample processing, and data analysis. Funding for this research was provided by the Marine Fisheries Advisory Board, The Charleston Harbor Project, and the

Department of Natural Resources. This paper is a portion of a

master’s thesis submitted to the University of Charleston, is contribution number 448 of the South Carolina Department of Natural Resources Marine Resources Division, and is contribution

number 2325 of the Virginia Institute of Marine Science.

LITERATURE CITED

ANDERSON, J. R., E. E. HARDY, J. T. ROACH, AND R. E. WITMER.

1976. A Land Use and Land Cover Classification System for

Use with Remote Sensor Data. U.S. Geological Survey Professional Paper 964, Reston, Virginia.

ARNOLD, C. L. AND C. J. GIBBONS. 1996. Impervious surface coverage. Journal of the American Planning Association 62:243–258.

BACA, B. J., J. DINGMAN, AND T. E. LANKFORD. 1988. Evaluation

of the impacts of dead-end canals and impoundments on an

estuarine ecosystem, p. 201–212. In W. L. Lyke and T. J. Hoban (eds.), Proceedings of the Symposium on Coastal Water

Resources, American Water Resources Association. Bethesda,

Maryland.

BIGGS, R. B., T. B. DEMOSS, M. M. CARTER, AND E. L. BEASLEY.

1989. Susceptibility of U.S. estuaries to pollution. Aquatic Sciences 1:189–207.

BILYARD, G. R. 1987. The value of benthic infauna in marine

pollution monitoring studies. Marine Pollution Bulletin 18:581–

585.

BROOM, M. J., J. M. DAVIES, B. HUTCHINGS, AND W. HALCROW.

1991. Environmental assessment of the effects of polluting

discharges: Stage I: Developing a post-facto baseline. Estuarine, Coastal and Shelf Science 33:71–87.

BROWN, R. G. 1988. Effects of precipitation and land use on

storm runoff. Water Resources Bulletin 24:421–426.

CHAPMAN, P. M., M. A. FARRELL, AND R. O. BRINKHURST. 1982.

Relative tolerances of selected aquatic oligochaetes to individual pollutants and environmental factors. Aquatic Toxicology 2:

47–67.

COHEN, J. E., C. SMALL, A. MELLINGER, J. GALLUP, AND J. SACHS.

1997. Estimates of coastal populations. Science 278:1211–1212.

CULLITON, T. J., M. A. WARREN, T. R. GOODSPEED, D. G. REMER,

C. M. BLACKWELL, AND J. J. I. MCDONOUGH. 1990. The Second

Report of a Coastal Trends Series: 50 Years of Population

Change along the Nation’s Coasts 1960–2010. U.S. Department of Commerce, National Oceanic and Atmospheric Administration, Rockville, Maryland.

DIAZ, R. J. 1980. Ecology of tidal freshwater and estuarine Tubificidae (Oligochaeta), p. 319–330. In R. O. Brinkhurst and

D. G. Cook (eds.), Aquatic Oligochaete Biology. Plenum

Press, New York.

DIAZ, R. J. 1992. Ecosystem assessment using estuarine and marine benthic community structure, p. 67–85. In G. A. Burton,

Jr. (ed.), Sediment Toxicology Assessment. Lewis Publishing,

Boca Raton, Florida.

DIAZ, R. J. AND R. ROSENBERG. 1995. Marine benthic hypoxia: A

review of its ecological effects and the behavioral responses

of benthic macrofauna. Oceanography and Marine Biology: An

Annual Review 33:245–305.

DODD, M. G. AND T. M. MURPHY. 1996. Management Recommendations for Colonial Waterbirds in the Charleston Harbor

Estuary. South Carolina Department of Natural Resources,

Charleston, South Carolina.

DRIVER, N. E. AND B. M. TROUTMAN. 1989. Regression models

for estimating urban storm-runoff quality and quantity in the

United States. Journal of Hydrology 109:221–236.

EDWARDS, S. 1989. Estimates of future demographic changes in

the coastal zone. Coastal Management 17:229–240.

ENGLE, V. D. AND J. K. SUMMERS. 1998. Determining the causes

of benthic condition. Environmental Monitoring and Assessment

51:381–397.

ENGLE, V. D., K. J. SUMMERS, AND G. R. GASTON. 1994. A benthic

index of environmental condition of Gulf of Mexico Estuaries. Estuaries 17:372–384.

FORTNER, A. R., M. SANDERS, AND S. W. LEMIRE. 1996. Polynuclear aromatic hydrocarbon and trace metal burdens in sediment and the oyster, Crassostrea virginica Gmelin, from twohigh salinity estuaries in South Carolina, p. 445–476. In F. J.

Vernberg, W. B. Vernberg, and T. Siewicki (eds.), Sustainable

Development in the Coastal Zone. University of South Carolina Press, Columbia, South Carolina.

FRETWELL, S. D. 1987. Food chain dynamics: The central theory

of ecology? Oikos 50:291–301.

FULTON, M. H., G. I. SCOTT, A. FORTNER, T. F. BIDLEMAN, AND

B. NGABE. 1993. The effects of urbanization on small high

salinity estuaries of the southeastern United States. Archives of

Environmental Contamination and Toxicology 25:476–484.

GASTON, G. R. 1985. Effects of hypoxia on macrobenthos of the

inner shelf off Cameron, Louisiana. Estuarine, Coastal and Shelf

Science 20:603–613.

GIERE, O. AND O. PFANNKUCKE. 1982. Biology and ecology of

marine oligochaeta. A review. Oceanography and Marine Biology:

an Annual Review 20:173–308.

GRASSLE, J. F. AND J. P. GRASSLE. 1974. Opportunistic life histories and genetic systems in marine benthic polychaetes. Journal of Marine Research 32:253–284.

HACKNEY, C. T., W. D. BURBANCK, AND O. P. HACKNEY. 1976. Biological and physical dynamics of a Georgia tidal creek. Chesapeake Science 17:271–280.

HOLLAND, A. F., G. H. M. RIEKERK, S. B. LERBERG, L. E. ZIMMERMAN, D. M. SANGER, T. D. MATTHEWS, G. I. SCOTT, M. H. FULTON, B. C. THOMPSON, J. W. DAUGOMAH, J. C. DEVANE, K. M.

BECK, AND A. R. DIAZ. 1996. The Tidal Creek Project, Interim

Report. Charleston Harbor Project. South Carolina Department of Natural Resources, Marine Research Division,

Charleston, South Carolina.

HORLICK, R. G. AND C. B. SUBRAHMANYAM. 1983. Macroinvertebrate infauna of a salt marsh tidal creek. Northeast Gulf Science

6:79–89.

HUNTER, J. AND D. R. ARTHUR. 1978. Some aspects of the ecology

of Peloscolex benedeni Udekem (Oligochaeta: Tubificidae) in

the Thames Estuary. Estuarine, Coastal and Shelf Science 6:197–

208.

KARR, J. R. 1981. Assessment of biotic integrity using fish communities. Fisheries 6:21–27.

KARR, J. R. 1991. Biological integrity: A long-neglected aspect of

water resource management. Ecological Applications 1:66–84.
852 S. B. Lerberg et al.

KARR, J. R. AND D. R. DUDLEY. 1981. Ecological perspective on

water quality goals. Environmental Management 5:55–68.

KUCKLICK, J. R., S. K. SIVERTSEN, M. SANDERS, AND G. I. SCOTT.

1997. Factors influencing polycyclic aromatic hydrocarbon

distributions in South Carolina estuarine sediments. Journal of

Experimental Marine Biology and Ecology 213:13–29.

LASALLE, M. W., M. C. LANDIN, AND J. G. SIMS. 1991. Evaluation

of the flora and fauna of a Spartina Alterniflora marsh established on dredged material in Winyah Bay, South Carolina.

Wetlands 11:191–208.

LERBERG, S. B. 1997. Effects of watershed development on macrobenthic communities in tidal creeks of the Charleston

Harbor area. Master’s Thesis. University of Charleston,

Charleston, South Carolina.

LEVIN, L. A. 1984. Life history and dispersal patterns in a dense

infaunal polychaete assemblage: Community structure and response to disturbance. Ecology 65:1185–1200.

LEVIN, L. A. AND D. V. HUGGETT. 1990. Implications of alternative reproductive modes for seasonality and demography in

an estuarine polychaete. Ecology 71:2191–2208.

LONG, E. R., L. J. FIELD, AND D. D. MACDONALD. 1998. Predicting

toxicity in marine sediments with numerical sediment quality

guidelines. Environmental Toxicology and Chemistry 17:714–727.

LONG, E. R., D. D. MACDONALD, S. L. SMITH, AND F. D. CALDER.

1995. Incidence of adverse biological effects within ranges of

chemical concentrations in marine and estuarine sediments.

Environmental Management 19:81–97.

MAXTED, J. R., S. B. WEISBERG, J. C. CHAILLOU, R. A. ESKIN, AND

F. W. KUTZ. 1997. The ecological condition of dead-end canals

of the Delaware and Maryland coastal bays. Estuaries 20:319–

327.

MCCALL, P. L. 1977. Community patterns and adaptive strategies

of the infaunal benthos of Long Island Sound. Journal of Marine Research 35:221–266.

MCCANN, L. D. AND L. A. LEVIN. 1989. Oligochaete influence on

settlement, growth and reproduction in a surface-deposit

feeding polychaete. Journal of Experimental Marine Biology and

Ecology 131:233–253.

MIDDLEDITCH, B. S., S. R. MISSLER, D. G. WARD, J. B. MCVEY, A.

BROWN, AND A. L. LAWERENCE. 1979. Maturation of penaeid

shrimp: Dietary fatty acids. Proceedings of the World Mariculture

Society 10:472–476.

MILLER, R. A. AND H. C. J. MATTRAW. 1982. Storm water runoff

quality from three land-use areas in South Florida. Water Resources Bulletin 18:513–519.

NATIONAL OCEANIC AND ATMOSPHERIC ADMINISTRATION NATIONAL STATUS AND TRENDS PROGRAM. 1991. Second Summary of

Data on Chemical Contaminants in Sediments. National Oceanic and Atmospheric Administration Technical Memo. NOS

ORCA 59. Office of Oceanography and Marine Assessment,

Rockville, Maryland.

NUMMEDAL, D., G. F. OERTEL, D. K. HUBBARD, AND A. C. HINE.

1977. Tidal inlet variability—Cape Hatteras to Cape Canaveral. Coastal Sediments: 5th Symposium of the Waterway. Port

Coastal and Ocean Division, American Society of Engineers,

New York.

OLSEN, R., N. H. CUTSHALL, AND I. L. LARSEN. 1982. Pollutantparticle associations and dynamics in coastal marine environments: A review. Marine Chemistry 11:501–533.

PEARSON, T. H. AND R. ROSENBERG. 1978. Macrobenthic succession in relation to organic enrichment and pollution of the

marine environment. Oceanography and Marine Biology: An Annual Review 16:229–311.

PFANNKUCHE, O. 1978. Abundance and lifecycle of littoral marine and brackish-water Tubificidae and Naididae (Oligochaeta), p. 103–111. In E. Naylor and R. G. Hartnoll (eds.),

Cyclic Phenomena in Marine Plants and Animals: Proceedings of the 13th European Marine Biology Symposium. Pergamon Press, New York.

PIMM, S. L. AND R. L. KITCHING. 1987. The determinants of food

chain lengths. OIKOS 50:302–307.

PLUMB, JR., R. H. 1981. Procedures for Handling and Chemical

Analysis of Sediment and Water Samples. Technical Report

EPA/CE-81–1. U.S. Environmental Protection Agency/Corps

of Engineers Technical Committee on Criteria for Dredged

and Filled Material. Environmental Laboratory, U.S. Army

Waterways Experiment Station, Vicksburg, Mississippi.

SANGER, D. M. 1998. Physical, chemical and biological environmental quality of tidal creeks and salt marshes in South Carolina estuaries. Ph.D. Dissertation, University of South Carolina, Columbia, South Carolina.

SANGER, D. M., A. F. HOLLAND, AND G. I. SCOTT. 1999a. Tidal

creek and salt marsh sediments in South Carolina coastal estuaries: I. Distribution of trace metals. Archives of Environmental

Contamination and Toxicology 37:445–457.

SANGER, D. M., A. F. HOLLAND, AND G. I. SCOTT. 1999b. Tidal

creek and salt marsh sediments in South Carolina coastal estuaries. II. Distribution of organic contaminants. Archives of

Environmental Contamination and Toxicology 37:458–471.

SARDA, R., I. VALIELA, AND K. FOREMAN. 1996. Decadal shifts in

a salt marsh macroinfaunal community in response to sustained long-term experimental nutrient enrichment. Journal

of Experimental Marine Biology and Ecology 205:63–81.

SAS INSTITUTE INC. 1989. SAS/STAT User’s Guide, Version 6,

4th edition. SAS Institute Inc., Cary, North Carolina.

SCHUELER, T. 1994. The importance of imperviousness. Watershed Protection Techniques 1:100–111.

SHENKER, J. M. AND J. M. DEAN. 1979. The utilization of an intertidal salt marsh creek by larval and juvenile fishes: Abundance, diversity and temporal variation. Estuaries 2:154–163.

TEAL, J. M. 1962. Energy flow in the salt marsh ecosystem of

Georgia. Ecology 43:614–624.

TIGER/LINE FILES. 1992. Technical documentation. The Bureau

of the Census. The Bureau, 1993, Washington, D.C.

VAN DEN BROEK, W. L. F. 1978. Dietary habits of fish populations

in the Lower Medway estuary. Journal of Fish Biology 3:645–654.

VAN DOLAH, R. F., J. F. HYLAND, A. F. HOLLAND, J. S. ROSEN, AND

T. R. SNOOTS. 1999. A benthic index of biological integrity for

assessing habitat quality in estuaries of the southeastern United States. Marine Environmental Research 48:269–283.

VERNBERG, F. J., W. B. VERNBERG, E. BLOOD, A. FORTNER, M. H.

FULTON, H. MCKELLAR, W. MICHENER, G. SCOTT, T. SIEWICKI,

AND K. EL FIGI. 1992. Impact of urbanization on high-salinity

estuaries in the southeastern united states. Netherlands Journal

of Sea Research 30:239–248.

WARWICK, R. M. AND K. R. CLARKE. 1991. A comparison of some

methods for analyzing changes in benthic community structure. Journal of Marine Biological Association of the United Kingdom 71:225–244.

WEINSTEIN, M. P., S. L. WEISS, AND M. F. WALTERS. 1980. Multiple

determinants of community structure in shallow marsh habitats, Cape Fear River Estuary, North Carolina, USA. Marine

Biology 58:227–243.

WEISBERG, S. B., A. F. FRITHSEN, A. F. HOLLAND, J. F. PAUL, K. J.

SOTT, J. K. SUMMERS, H. T. WILSON, R. VALENTE, D. G. HEIMBUCH, J. GERRISEN, S. C. SCHIMMEL, AND R. W. LATIMER. 1990.

EMAP-Estuaries Virginian Province 1990 Demonstration Project Report. EPA-600/R-92/100. U.S. Environmental Protection Agency, Environmental Research Laboratory, Narragansett, Rhode Island.

WEISBERG, S. B., J. A. RANASINGHE, D. M. DAUER, L. SCHAFFNER,

AND R. J. DIAZ. 1997. An estuarine benthic index of biotic

integrity. Estuaries 20:149–158.

WENNER, C. 1992. Red Drum: Natural History and Fishing Techniques in South Carolina. Marine Resources Division, South

Carolina Wildlife and Marine Resources Department,

Charleston, South Carolina.

WENNER, E. L. AND H. R. BEATTY. 1993. Utilization of shallow
Watershed Development in Tidal Creeks 853

estuarine habitats in South Carolina, U.S.A., by postlarval and

juvenile stages of Penaeus spp. (Decapoda: Penaeidae). Journal

of Crustacean Biology 13:280–295.

WENNER, E., A. F. HOLLAND, AND D. SANGER. 1999. Assessing

short-term variability in dissolved oxygen and other water

quality variables in shallow estuarine habitats, p. 802–806. In

Ocean Community Conference 1998 Proceedings, Volume 2,

November 16–19, 1998. Marine Technology Society, Washington, D.C.

WIEGERT, R. G. AND B. J. FREEMAN. 1990. Tidal salt marshes of

the southeast Atlantic Coast: A community profile. United

States Fish and Wildlife Service Biological Report 85:1–67.

WILTSE, W. I., K. H. FOREMAN, J. M. TEAL, AND I. VALIELA. 1984.

Effects of predators and food resources on the macrobenthos

of salt marsh creeks. Journal of Marine Research 42:923–942.

Received for consideration, May 26, 1999

Accepted for publication, August 1, 2000