1997-1998 Final Report ENVIRONMENTAL QUALITY OF WILMINGTON AND NEW HANOVER COUNTY WATERSHEDS
1997-1998
by
Michael A. Mallin, Lawrence B. Cahoon, John J. Manock, James F. Merritt, Martin H. Posey, Troy D. Alphin, Douglas C. Parsons and Tracey L. Wheeler
CMSR Report 98-03 Center for Marine Science Research University of North Carolina at Wilmington Wilmington, N.C. 28403
November, 1998
Funded by: City of Wilmington
New Hanover County
Northeast New Hanover Conservancy North Carolina State University
Executive Summary
This report represents combined results of Year 5 of the New Hanover County
Tidal Creeks Project and Year 1 of the Wilmington Watersheds Project. Water quality data are presented from a watershed perspective, regardless of political boundaries. The combined programs involved eleven watersheds and 46 sampling stations.
Barnards Creek – There were no major surface water quality problems at the stations
sampled in the Barnards Creek watershed. Incoming and outflowing water at the wet detention pond on Echo Farms Golf course were sampled, with nitrate, ammonium and total phosphorus (TP) all significantly higher in the stream exiting the course. Even so, the concentrations of those nutrients were low in comparison to other area golf courses
sampled, possibly because of nutrient uptake in a natural wetland through which the
outfall stream passes before leaving the course. Concentrations of metals in sediments in the pond exit stream were somewhat greater than in the incoming stream, although they were not at harmful levels. Metals in sediments of a tributary stream near Titanium Road were very high, and at biologically harmful levels for cadmium, chromium and
zinc, indicating a metals source upstream of this wooded site.
Bradley Creek – Turbidity was a major problem in this watershed, particularly downstream of construction areas in the north branch of Bradley Creek. Low dissolved oxygen was also a problem in brackish waters of the creek, especially in summer.
Elevated nitrogen and phosphorus levels enter the creek in both the north and south
branches, leading to spring and summer algal blooms, particularly in the south branch near Wrightsville Avenue. Fecal coliform pollution remains problematic in Bradley Creek, especially in the upper branches, with the largest inputs in the south branch.
Burnt Mill Creek – A sampling station on Burnt Mill creek at Princess Place had
substandard dissolved oxygen during 50% of the sampling trips. This station also had poor microbiological water quality, exceeding the standard for human contact in six of 10 samples. The effectiveness of Ann McCrary wet detention pond on Randall Parkway as a pollution control device was mixed. The pond did significantly reduce turbidity and
fecal coliform levels. However, no decreases in nutrients were realized, possibly
because of suburban inputs along the middle and lower shore of the pond. Levels of arsenic, iron, and nickel were significantly higher downstream of the pond, although they were not at harmful levels.
Futch Creek – Futch Creek maintained good microbial water quality, as it has since
channel dredging at the mouth occurred in 1995 and 1996. Algal blooms were not a problem in 1997-1998. Elevated nitrate loading continued to occur at stations in the upper south branch. An investigation of nutrient concentrations in springs indicated that nitrate-rich groundwater was entering the creek in this area, the original source of which
is currently unknown.
Greenfield Lake – Two tributaries of Greenfield Lake (near Lake Branch Drive and Lakeshore Commons Apartments) had low dissolved oxygen problems on 10 of 12 occasions. Both of these tributaries also had frequent high fecal coliform counts, above
the state standard for human contact waters. The stream near Lakeshore Commons
also maintained high nitrate and phosphate concentrations. Stations within the lake all experienced algal blooms at times, consisting mainly of green and blue-green species. Generally, nutrient loading and low dissolved oxygen problems were most severe at a
station located in the south end that receives several urban and suburban inputs. A
large regional wet detention pond on the tributary Silver Stream did a very good job of reducing pollutant loads to the lake from this drainage. Statistically significant reductions in nitrate, ammonium, total nitrogen (TN), orthophosphate, TP, aluminum, chromium and iron were all realized. The design of this pond consists of two
interconnected basins containing large amounts of diverse aquatic vegetation, with most
inputs directed into the upper basin. This could serve as a potential model for future pond design. Hewletts Creek – This creek receives high nutrient loading in its three upper branches,
with consequent algal blooms especially in the south branch near Pine Grove Road.
The middle branch had the highest nutrient concentrations, largely derived from two golf courses. Results of nutrient addition bioassay experiments indicated that nitrogen stimulated phytoplankton growth in the lower creek in both spring and summer, and in the south branch during summer.
Howe Creek – Lower Howe Creek near the ICW maintained generally good water quality. An analysis of pesticide runoff did not detect organic pesticides in a receiving area (Graham Pond) of runoff from a Landfall golf course. Nutrient loading and algal bloom problems were most severe near and upstream of the Graham Pond input area.
There is a blowdown pipe discharging potable water from the city system into the creek
near Graham Pond. This water contains high nutrient levels and may contribute to algal bloom formation at this location. Turbidity and low dissolved oxygen were problems in the upper areas of Howe Creek. Land clearing and construction activities in the upper watershed are undoubtedly the cause of non-point source pollution to the upper creek.
An analysis of flushing rates found that during dry periods flushing rates in Howe Creek
were equivalent to those of Futch Creek after dredging. During wet periods the flushing rate drops significantly, probably because of formation of a freshwater lens impeding mixing. The amount of dredging required to alter the creek to increase mixing would be prohibitively expensive. Efforts to reduce stormwater inputs to the creek (buffer zones,
large wet detention ponds, use of natural wetlands, etc.) would be more cost effective
and directly treat the causes of both the pollution and runoff problems. Pages Creek – This creek maintained generally good water quality during 1997-1998. Nutrient loading and phytoplankton growth was low to moderate, even at the most
anthropogenically-impacted stations. The only problems were periodic hypoxia is
summer at the stations draining Bayshore Drive. Smith Creek – Turbidity was a problem at all three stations sampled in this watershed. Runoff from highway construction was a likely contributing factor to this. Nutrient and
chlorophyll a concentrations were unremarkable. However, fecal coliform bacterial
counts periodically exceeded the state standard for human contact waters at all three sites. Upper and Lower Cape Fear Watersheds – Water quality at the one station on the
stream draining the Upper Cape Fear watershed (behind the Wilmington police station)
was generally good except for very high nitrate concentrations, the source of which is presently not known. Sediment concentrations of arsenic and lead were high here as well, with lead concentrations in the range of those occasionally harmful to aquatic life.
Pfiesteria survey – Thirteen stations throughout the tidal creek complex were sampled in summer 1997 for the presence of the toxic dinoflagellate Pfiesteria piscicida. This organism was found only during June, July and August, in very few locations. The stations where it occurred were stations showing periodic signs of eutrophication,
including the north and south branches of Bradley Creek at Wrightsville Avenue, the
south branch of Hewletts Creek at Pine Grove Road, the south branch of Futch Creek, and the two Bayshore Drive stations on Pages Creek. Pfiesteria counts were positively correlated with chlorophyll a, total phosphorus, and nitrate, and negatively correlated with salinity. We emphasize that the Pfiesteria collected were in non-toxic stages, and
we are unaware of any fish kills in these creeks related to Pfiesteria at present.
However, since Pfiesteria is present, and there are appropriate environmental conditions in these tidal creeks to support its growth, the possibility remains that this organism may cause fish kills locally at some point in time.
Benthic Studies – The fall 1997 survey found 12 bivalve taxa in the tidal creek system,
the most common species being Tagelus plebius. There were no differences in overall bivalve abundance among the creeks, but diversity was much lower in Howe Creek then the other creeks. This pattern has been noted in several earlier surveys as well. During the past year a number of habitat utilization tests and bivalve growth experiments were
initiated in the tidal creeks, preliminary results of which are reported within.
Wet Detention Pond Studies – A key finding from ongoing wet detention pond studies was that the phosphorus content in sediments increases with pond age until approximately ten years, after which it remains relatively constant. This indicates that
older ponds may export phosphorus to downstream receiving water bodies. However,
no relationship was found between biomass of cyanobacteria (blue-green algae) and pond water or sediment phosphorus concentrations. Thus, growth of these nitrogen-fixing species is likely controlled by a complex of factors, including the role of aquatic macrophyte species.
Table of Contents
1.0 Introduction 1
1.1 Methods 1 2.0 Barnards Creek 4
3.0 Bradley Creek 8 4.0 Burnt Mill Creek 12
5.0 Futch Creek 15
6.0 Greenfield Lake 19 7.0 Hewletts Creek 23
8.0 Howe Creek 28 8.1 Tidal Flushing of Howe Creek 31
9.0 Pages Creek 34
10.0 Smith Creek 36 11.0 Upper and Lower Cape Fear 38
12.0 Pfiesteria Survey in the Tidal Creeks 42 13.0 Tidal Creek Benthic Community 44
14.0 Tidal Creek Pesticide Study 52
15.0 Stormwater Detention Pond Study 61 16.0 References Cited 72
17.0 Acknowledgments 75 18.0 Appendix A 76
1.0 Introduction Since 1993, scientists at the UNC Wilmington Center for Marine Science
Research have been studying five tidal creeks in New Hanover County. This project,
funded by New Hanover County, the Northeast New Hanover Conservancy, and UNCW, yielded a comprehensive report detailing important findings from 1993-1997, and produced a set of management recommendations for improving creek water quality (Mallin et al. 1998). In the present report we present results of continuing studies from
1997-1998 in the tidal creek complex (Fig. 1.1).
. Additionally, in October 1997 the Center for Marine Science Research began a project (funded by the City of Wilmington Engineering Department) with the goal of assessing water quality in Wilmington City watersheds under base flow conditions.
Additionally, certain sites were analyzed for sediment heavy metals concentrations
(EPA Priority Pollutants). The water quality data within is presented from a watershed perspective. Some of the watersheds cross political boundaries (i.e. parts of the same watershed may lie in
the County but not the City). Bradley and Hewletts Creeks are examples. Water quality
parameters analyzed in the tidal creeks include water temperature, pH, dissolved oxygen, salinity/conductivity, turbidity, nitrate, ammonium, orthophosphate, chlorophyll
a, and in selected creeks fecal coliform bacteria. Similar analyses were carried out in the City watersheds with the addition of total nitrogen, total phosphorus, and suspended
solids. The current results of an ongoing (five year) analysis of the tidal creek benthic
macrofauna are also presented. Also within this report are results of related studies. These include an analysis of the presence of the dinoflagellate Pfiesteria piscicida relative to water quality, an
assessment of herbicide runoff into Howe Creek, an estimation of flushing time for
Howe Creek, and another is part of a continuing study on the functioning of wet detention ponds. 1.1 Methods
Field parameters were measured at each site using either a YSI 6920 Multiparameter Water Quality Meter and Probe (sonde) or a Solomat 803PS Multiparameter sonde coupled with a Solomat 803 datalogger. Individual probes within the instruments measured water temperature, pH, dissolved oxygen, turbidity, salinity,
and conductivity. YSI Model 85 and 55 dissolved oxygen meters were also used on
occasion. The instruments were calibrated prior to and after each sampling trip to ensure accurate measurements. For the five tidal creeks, water samples were collected monthly, at or near high
tide. For nitrate+nitrite (hereafter referred to as nitrate) and orthophosphate
assessment, three replicate acid-washed 125 mL bottles were placed ca. 10 cm below the surface, filled, capped, and stored on ice until processing. In the laboratory the
triplicate samples were filtered simultaneously through 1.0 µM glass fiber filters using a
manifold with three funnels. The pooled filtrate was stored frozen until analysis.
Nitrate+nitrite and orthophosphate were analyzed using a Technicon AutoAnalyzer
following EPA protocols. Samples for ammonium were collected in duplicate, field-preserved with phenol, stored on ice, and analyzed in the laboratory according to the methods of Parsons et al. (1984). For the tidal creeks, ammonium was sampled only
during spring and summer, the principal phytoplankton growing season. Fecal coliform
samples were collected by filling pre-autoclaved containers ca. 10 cm below the surface facing into the stream. Samples were stored on ice until processing (< 6 hr). Fecal coliform concentrations were determined using a membrane filtration (mFC) method (APHA 1995).
The analytical method used to measure chlorophyll a is described in Welschmeyer (1994) and US EPA (1997). Chlorophyll a concentrations were determined directly from the 1.0 micron glass fiber filters used for filtering samples for nitrate+nitrite and orthophosphate analyses. All filters were wrapped individually in
aluminum foil, placed in an airtight container containing drierite, and stored in the
freezer. During the analytical process, the glass filters were separately immersed in 10 ml of a 90% acetone solution for 24 hours. The acetone extracts the chlorophyll a from the glass filters into solution. Each solution was then analyzed for chlorophyll a concentrations using a Turner AU-10 fluorometer. This method uses an optimal
combination of excitation and emission bandwidths which reduces the errors inherent in
the acidification technique. Samples were collected monthly within the Wilmington City watersheds from October 1997 through July 1998. Field measurements were taken as indicated above.
Nutrients (nitrate, ammonium, total Kjeldahl nitrogen (TKN), total nitrogen (TN),
orthophosphate, and total phosphorus (TP)) and total suspended solids (TSS) were analyzed by a state-certified contract laboratory using EPA and APHA techniques. Chlorophyll a was run at UNCW-CMSR as above. Sediment analyses for EPA priority pollutants were run in triplicate from 12 sites in the Wilmington watershed system. EPA
methods were used for digestion (Method 3050B) and ICP analysis (Method 200.7).
Guidelines for what various sediment metals concentrations represent in terms of potential harm to aquatic life are presented in Appendix A. For three detention ponds (Ann McCrary Pond on Burnt Mill Creek, Silver Stream
Pond in the Greenfield Lake watershed, and the main pond on Echo Farms Golf Course
in the Barnards Creek watershed) we were able to obtain data from input (control) and outfall stations. We used these data to test for statistically significant differences in pollutant concentrations between pond input and output stations. The data were first tested for normality using the Shapiro-Wilk test. Normally distributed data parameters
were tested using the paired-difference t-test, and non-normally distributed data
parameters were tested using the Wilcoxon Signed Rank test. Statistical analyses were conducted using SAS (Schlotzhauer and Littell 1987). Sediment metal concentrations for input and outflow pond stations were compared using t-tests.
2.0 Barnards Creek The BNC-TR site in Barnard’s Creek watershed drains a wooded area and was
considered a control site for nutrients and physical parameters (Fig.2.1). The BNC-CB
site is near Carolina Beach Road and drains an area hosting construction activities. Inorganic nutrients (nitrate, ammonium, and orthophosphate) and turbidity were all considerably larger at BNC-CB than BNC-TR (Table 2.1). Total N and P did not differ between the two sites, reflecting the larger organic nutrient quantity found in the wooded
stream site.
The Barnard’s Creek watershed hosts the Echo Farms Country Club. Much of the course drainage enters a large (1.75 acre) pond, which discharges through a wooded wetland. We sampled a pond input station draining a suburban area (BNC-EF)
and the pond outflow exiting the riparian wetland (BNC-AW). A primary goal was to
assess the efficacy of golf course nutrient removal by this pond. Golf course loading led to significant increases in nitrate, ammonium, and TP, but there were no significant differences in the other pollutant parameters tested (Table 2.1). There were no apparent input-output differences in nutrient concentration during fall and winter of
1997-1998, but from late spring through summer 1998 there were considerable
increases in nutrient loading from the golf course (Fig.2.2), possibly a result of increased course fertilization during the growing season (Walker and Branham 1992). We do point out that overall concentrations of inorganic nutrients in the output stream were low in comparison with output from other area golf courses (Mallin and Wheeler
1998). The design of a wet detention pond followed by a natural wetland area does a
good job of keeping nutrient concentrations low in the course outfall. Table 2.1. Mean and standard deviation of water quality parameters in Barnard’s Creek watershed, including a comparison of pollutant concentrations in input and output
waters of wet detention pond on Echo Farms Country Club, October 1997-July 1998.
_____________________________________________________________________ Parameter BNC-TR BNC-CB BNC-EF BNC-AW _____________________________________________________________________
DO (mg/L) 6.8 (1.4) 6.7 (1.8) 6.4 (1.0) 6.4 (1.2)
Turbidity (NTU) 6.0 (7.2) 15.7 (16.3) 18.1 (22.6) 14.8 (18.7) TSS (mg/L) 2.5 (1.6) 2.2 (0.9) 3.0 (1.7) 5.0 (5.1) Nitrate (mg/L) 0.013 (0.013) 0.073 (0.036) 0.024 (0.021) 0.059 (0.044)* Ammon. (mg/L) 0.018 (0.010) 0.054 (0.033) 0.019 (0.012) 0.030 (0.018)*
TN (mg/L) 0.645 (0.436) 0.655 (0.231) 0.597 (0.287) 0.611 (0.255)
Phosphate (mg/L) 0.021 (0.013) 0.027 (0.015) 0.035 (0.023) 0.041 (0.026) TP (mg/L) 0.031 (0.016) 0.028 (0.013) 0.044 (0.018) 0.060 (0.024)*
Chlor. a (µg/L) 1.0 (1.3) 1.8 (0.9) 1.6 (1.3) 1.9 (1.6)
Fec. col.(/100 mL) 63 78 44 27
_____________________________________________________________________ * Indicates significant difference between input and output concentration at p<0.05 **Indicates significant difference between input and output concentration at p<0.01
Sediment metals concentrations were assessed at the input and outflow stations
of the wet detention pond at Echo Farms as well as the forested station at Titanium road
(Fig. 2.1). The data indicated that sediment metals were very high at BNC-TR, in many cases the highest in the entire system (Table 2.2). Both cadmium and chromium were at concentrations in which harmful effects to aquatic life occasionally occur, and zinc
was close to that range (Appendix A). This indicated a source upstream of that station
which is undergoing serious metals leaching into this stream. It is likely that a dumpsite upstream is the cause of this metals pollution. The wet detention pond analysis showed that every metal parameter was greater in the pond outflow than the input stream, but due to high variability among the replicate samples none of the differences were
statistically significant (Table 2.2). Golf Course chemicals and parking lot runoff are
likely contributing sources of these elevated metals. Table 2.2. Mean and (standard deviation) of sediment metals concentrations in Barnard’s Creek watershed, including a comparison of pollutant concentrations in input
and output waters of wet detention pond on Echo Farms Golf Course.
_____________________________________________________________________ Parameter BNC-TR BNC-EF BNC-AW _____________________________________________________________________
Al (mg/kg) 5620.0 (340.4) 633.0 (181.6) 5966.7(4018.5)
As (mg/kg) 0.726 (0.626) 0.017 (0.003) 0.585 (0.587) Cd (mg/kg) 1.802 (0.892) 0.076 (0.021) 0.597 (0.380) Cr (mg/kg) 320.7 (178.2) 0.80 (0.20) 6.59 (4.05) Cu (mg/kg) 28.07 (15.72) 0.94 (0.75) 14.27 (13.72)
Fe (mg/kg) 5126.7 (975.0) 484.0 (98.8) 4556.7 (3477.4)
Pb (mg/kg) 11.94 (2.30) 2.09 (0.50) 7.33 (6.65) Hg (mg/kg) 0.013 (0.001) 0.004 (0.002) 0.005 (0.002) Ni (mg/kg) 2.14 (0.45) 0.37 (0.10) 1.81 (1.30) Zn (mg/kg) 142.6 (84.6) 16.7 (10.3) 21.3 (18.1)
_____________________________________________________________________
* Indicates significant difference between input and output concentration at p<0.05 **Indicates significant difference between input and output concentration at p<0.01
3.0 Bradley Creek The Bradley Creek watershed is of particular current interest as a principal
location for future Clean Water Trust Fund mitigation activities, including the purchase
of Airlie Gardens by the County. This creek is one of the most polluted in New Hanover County, particularly by fecal coliform bacteria (Mallin et al. 1998). Seven stations were sampled in the past year, in both fresh and brackish waters (Fig. 3.1).
Turbidity has become a major problem in Bradley Creek (Table 3.1). Average
turbidity values exceeded the North Carolina standard for tidal waters at both BC-SB (which hosted a number of algal blooms) and BC-NB (which was impacted by upstream construction on Eastwood Road). BC-NBU, near Eastwood Road, was particularly impacted by road construction and had average turbidities exceeding the state standard
for fresh water (Table 3.1). BC-CA, on College Acres Road, periodically had high
turbidity due to commercial construction upstream. In June, July and August the downstream brackish stations periodically had hypoxic dissolved oxygen conditions. Sediment loading and low dissolved oxygen are environmental factors which can affect habitat and impair the fishery nursery function of Bradley Creek.
Table 3.1 Physical parameters at Bradley Creek sampling stations, 1997-1998. Data as mean (SD) / range. _____________________________________________________________________ Station Salinity Turbidity (NTU) Dissolved oxygen (mg/L)
_____________________________________________________________________
BC-76 28.5 (5.0) 13.7 (9.3) 6.7 (1.7) 17.7-35.6 3.5-24.5 3.2-8.7
BC-SB 4.6 (7.4) 26.8 (12.3) 7.5 (1.5)
0.2-26.3 16.8-43.2 4.7-9.8 BC-SBU 0.2 (0.2) 4.7 (2.1) 6.7 (1.5) 0.1-0.9 1.8-6.6 5.0-9.7
BC-NB 13.9 (11.6) 26.0 (16.1) 6.8 (1.9) 0.6-34.0 10.1-42.2 2.5-9.5 BC-NBU 0.1 (0.0) 66.9 (41.0) 7.2 (0.5)
0.1-0.2 15.5-100.0 6.4-8.2
BC-CR 0.1 (0.0) 3.1 (1.4) 7.0 (0.7) 0.0-0.1 1.5-5.0 6.0-7.8
BC-CA 0.1 (0.1) 38.3 (39.2) 6.4 (1.6)
0.0-0.2 6.2-121.3 3.9-8.3 _____________________________________________________________________ Nitrate concentrations were highest at stations BC-CR (draining a suburban
residential area along Clear Run Drive) and BC-SBU (upper south branch) and BC-CA
(draining apartment complexes and construction areas upstream). Particularly high orthophosphate levels were found at BC-CA, with somewhat elevated orthophosphate levels at both BC-SB and BC-SBU (Table 3.2). Ammonium was also elevated at BC-
CA. The nutrient loads led to particularly large phytoplankton blooms at BC-SB, and
lesser blooms at BC-NB (Table 3.2). Both of these stations have hosted dense blooms in the past, especially BC-SB (Mallin et al. 1998). Median inorganic molar N/P ratios were 10.1 for BC-76, 65.7 for BC-SB, and 86.3 for BC-NB. This indicates that phytoplankton growth at BC-76 is likely nitrogen-limited, but phosphorus limits growth in
the two tributary stations. However, ratios were greatest in early spring and lowest in
summer, indicating that nitrogen may become limiting during mid-to-late summer at all stations. Table 3.2 Nutrient and chlorophyll a data at Bradley Creek sampling stations, 1997-
1998. Data as mean (SD) / range, nutrients in mg/L, chlorophyll a as µg/L. _____________________________________________________________________ Station Nitrate Ammonium Orthophosphate Chlorophyll a _____________________________________________________________________
BC-76 0.013 (0.016) 0.018 (0.008) 0.006 (0.004) 2.5 (1.8) 0.001-0.061 0.007-0.029 0.002-0.014 0.4-5.7 BC-SB 0.073 (0.043) 0.046 (0.028) 0.009 (0.009) 19.9 (33.3)
0.015-0.146 0.010-0.087 0.002-0.034 1.6-108.9
BC-SBU 0.205 (0.221) NA 0.008 (0.007) 1.8 (1.6) 0.062-0.790 0.003-0.025 0.2-5.9
BC-NB 0.084 (0.071) 0.044 (0.045) 0.007 (0.008) 9.9 (11.7)
0.008-0.209 0.012-0.123 0.005-0.030 0.2-40.7 BC-NBU 0.121 (0.023) NA 0.006 (0.009) 2.2 (2.7) 0.086-0.159 0.005-0.027 0.3-7.4
BC-CR 0.236 (0.086) NA 0.007 (0.007) 0.7 90.4) 0.028-0.312 0.002-0.020 0.1-1.3 BC-CA 0.173 (0.141) 0.140 (0.053) 0.073 (0.067) 6.5 (9.3)
0.005-0.470 0.080-0.240 0.010-0.210 1.0-31.6
_____________________________________________________________________ Fecal coliform concentrations were generally high in this watershed (Table 3.3). In particular the south branch maintained geometric mean concentrations exceeding the
200 CFU/100 mL standard for human contact waters, and the north branch was
elevated to a somewhat lesser extent. The suburban feeder station maintained generally low fecal coliform concentrations but concentrations at BC-CA were much higher (Table 3.3). Both of these stations drain into the north branch of Bradley Creek (Fig. 3.1). Bradley Creek watershed is the most highly developed watershed in New
Hanover County with the greatest percent impervious surface areal coverage, a critical
factor exacerbating fecal coliform pollutant loading (Mallin et al. 1998).
Table 3.3 Fecal coliform abundance (geometric mean) at Bradley Creek sampling stations, August 1997-July 1998.
_____________________________________________________________________
Station BC-76 BC-SB BC-SBU BC-NB BC-NBU BC-CR BC-CA GEOMEAN 9 417 278 197 182 25 143 _____________________________________________________________________
Sediment metals concentrations were evaluated at two stations (Table 3.4). In all cases metals concentrations at BC-CA were greater (sometimes substantially so) than concentrations at BC-CR. BC-CA drains an area which has been subject to a considerable amount of construction activity in recent years, as well as established
apartment complexes, while BC-CR drains a well-established suburban neighborhood.
Table 3.4. Mean and standard deviation of sediment metals concentrations in two stations in the Bradley Creek watershed. _____________________________________________________________________
Parameter (mg/kg) BC-CA BC-CR
_____________________________________________________________________ Al 2986.7 (638.8) 364.0 (141.2) As 0.203 (0.170) 0.018 (0.003)
Cd 0.179 (0.126) 0.053 (0.010)
Cr 1.93 (1.26) 0.41 (0.10) Cu 1.34 (1.28) 0.33 (0.07) Fe 968.3 (212.3) 105.7 (44.1) Pb 6.18 (4.85) 0.60 (0.0.53)
Hg 0.006 (0.003) 0.004 (0.001)
Ni 0.59 (0.36) 0.20 (0.04) Zn 5.65 (4.55) 2.66 (0.45) _____________________________________________________________________
4.0 Burnt Mill Creek The Burnt Mill Creek watershed was sampled just upstream of Ann McCrary
Pond on Randall Parkway (AP1), along shore at mid-pond (AP2), about 40 m
downstream of the pond outfall (AP3), and in the creek from the bridge at Princess Place (BMC-PP - Fig. 4.1). Ann McCrary Pond is a large (28.8 acres) regional wet detention pond draining 1,785 acres, with active apartment construction at the upper end near AP1. The pond itself usually maintains a thick growth of submersed aquatic
vegetation, particularly Hydrilla verticillata, Egeria densa, Alternanthera philoxeroides,
Ceratophyllum demersum and Valliseneria americana. A survey in late summer 1998 indicated that approximately 70% of the pond area was vegetated. There have been efforts to control this growth, including addition of triploid grass carp as grazers. Our survey also found that this pond is host to Lilaeopsis carolinensis, which is a threatened
plant species in North Carolina.
The Princess Place location experienced some water quality problems during the sample period. Dissolved oxygen was substandard on 5 of 10 sampling trips. The most important issue, from a public health perspective, was the excessive fecal coliform
counts, which maintained a geometric mean well in excess of the state standard for
human contact waters (Table 4.1). Fecal coliform counts were greater than 200 CFU/100 mL in 6 of 10 samples. Turbidity and fecal coliform concentrations entering Ann McCrary Pond were very
high (Table 4.1), possibly a result of the extensive construction and land-clearing
activities during this study. The efficiency of this pond as a pollutant removal device was mixed. Turbidity and fecal coliforms were significantly reduced (Table 4.1). However, there was no significant difference in removal of total suspended solids or the other nutrient parameters. This may be due in part to increased phytoplankton growth
in the pond proper, some of which is then carried over the dam to the outfall site (Table
4.1). It is also likely that inputs of nutrients, especially nitrate, have entered the pond from suburban drainage stream midway down the pond across from our AP2 site. Also, intensive waterfowl use of the pond, particularly at a tributary near the outfall, may have contributed to nutrient loading in the pond and along its shoreline. Peak nitrate, TN, and
TP concentrations were in mid-pond. Dissolved oxygen significantly increased through
the pond, probably because of in-pond photosynthesis and aeration by passage over the final dam at the outfall.
Table 4.1. Mean and (standard deviation) of water quality parameters in Burnt Mill Creek watershed. Fecal coliforms given as geometric mean. _____________________________________________________________________
Parameter BMC-PP BMC-AP1 BMC-AP2 BMC-AP3
_____________________________________________________________________ DO (mg/L) 5.2 (1.5) 6.5 (1.1) 7.3 (1.4) 8.9 (1.1)** Turbidity (NTU) 15.8 (15.2) 33.1 (30.9) 24.0 (29.5) 15.0 (20.7)* TSS (mg/L) 9.5 (9.2) 5.7 (7.2) 5.8 (3.9) 4.7 (3.7)
Nitrate (mg/L) 0.221 (0.110) 0.143 (0.061) 0.229 (0.172) 0.209 (0.143)
Ammonium (mg/L) 0.135 (0.066) 0.094 (0.025) 0.088 (0.105) 0.067 (0.073) TN (mg/L) 1.124 (0.690) 0.716 (0.162) 0.879 (0.218) 0.832 (0.218) Phosphate (mg/L) 0.051 (0.030) 0.041 (0.025) 0.038 (0.039) 0.025 (0.029) TP (mg/L) 0.080 (0.028) 0.048 (0.020) 0.051 (0.033) 0.047 (0.026)
N/P molar ratio 17.6 14.6 29.1 31.3
Fec. col. (/100 mL) 295 303 108 39*
Chlor. a (µg/L) 5.6 (3.2) 2.8 (3.10) 10.0 (8.4) 8.2 (8.3) _____________________________________________________________________
* Indicates statistically significant difference between AP1 and AP3 at p<0.05
**Indicates statistically significant difference between AP1 and AP3 at p<0.01 Sediment metals were analyzed from the input and outflow stations of Ann McCrary Pond (Table 4.2). Concentrations were generally low entering the pond, but
tended to increase at the outflow station. Increases in arsenic, iron and nickel were all
statistically significant. However, metals at BMC-AP3 were still comparatively low. Table 4.2. Mean and standard deviation of sediment metals concentrations in input and output waters of Ann McCrary wet detention pond on Randall Parkway.
_____________________________________________________________________
Parameter (mg/kg) BMC-AP1 BMC-AP3 _____________________________________________________________________ Al 395.0 (120.8) 790.3 (176.2) As 0.022 (0.009) 0.130 (0.036)*
Cd 0.051 (0.006) 0.093 (0.016)
Cr 0.97 (0.06) 1.01 (0.41) Cu 0.45 (0.08) 0.78 (0.14) Fe 180.3 (41.0) 773.0 (86.5)* Pb 1.46 (0.22) 1.45 (0.51)
Hg 0.003 (0.000) 0.004 (0.002)
Ni 0.21 (0.05) 0.53 (0.12)* Zn 3.71 (0.44) 3.81 (1.86) _____________________________________________________________________ * Indicates significant difference between input and output concentration at p<0.05
**Indicates significant difference between input and output concentration at p<0.01
5.0 Futch Creek During 1995 and 1996 two channels were dredged in the mouth of Futch Creek
(Fig. 5.1) to improve circulation and hopefully reduce fecal coliform bacterial
concentrations. There was a statistically significant increase in salinity in the creek in the months following dredging (Mallin et al. 1996). Salinity continued to remain relatively high even in the upper portions of the creek in 1997-1998, demonstrating the continued effectiveness of the dredging operation (Table 5.1). None of the creek
stations had average turbidity exceeding the state standard of 25 NTU, although during
June, six of the eight stations exceeded the turbidity standard. There were incidences of substandard dissolved oxygen at all of the stations, with the most incidences (three) at FC-17 and FC-20.
Table 5.1 Physical parameters at Futch Creek sampling stations, 1997-1998. Data as
mean (SD) / range. _____________________________________________________________________ Station Salinity Turbidity (NTU) Dissolved oxygen (mg/L) _____________________________________________________________________
FC-2 31.7 (2.9) 10.4 (6.4) 8.2 (2.3) 25.7-36.1 1.8-19.6 4.7-10.9 FC-4 31.2 (3.1) 9.1 (5.9) 8.1 (2.4)
24.6-35.8 1.3-19.8 4.2-10.9
FC-6 30.5 (3.5) 11.1 (8.5) 8.0 (2.5) 23.3-35.6 1.6-27.6 4.1-11.3
FC-8 27.7 (5.3) 10.4 (7.0) 7.4 (2.6)
15.4-33.3 1.5-27.1 3.1-10.4 FC-13 23.7 (5.5) 14.4 (9.6) 7.5 (2.4) 12.9-31.8 2.6-36.5 2.7-10.0
FC-17 19.5 (7.5) 15.8 (8.9) 6.8 (2.8) 7.0-30.0 1.8-26.6 2.0-9.9 FC-20 16.8 (10.8) 13.4 (10.8) 6.7 (2.6)
0.7-30.5 1.4-42.0 1.9-9.7
FOY 24.8 (6.2) 10.2 (6.9) 7.2 (2.8) 13.6-32.5 1.5-24.1 2.7-10.4 _____________________________________________________________________
Nutrient concentrations in Futch Creek were generally low, with the exception of periodic nitrate pulses in the upper stations FC-17 and FC-20 (Table 5.2). The source of these pulses has been investigated (see below). The creek was free from algal blooms during our sampling visits (Table 5.2), even in the upper stations. Computed
median inorganic N/P molar ratios were 7.2 for FC-4 (indicating N limitation of
phytoplankton growth), but 37.3 and 29.3 for FC-17 and FOY, respectively (indicating P limitation of phytoplankton growth). N/P ratios in FOY decreased considerably in summer, possibly indicating a change to N limitation, but were at or above the Redfield
Ratio (16) in both spring and summer at FC-17. The reason for the elevated nitrogen at
this station is discussed below. Table 5.2 Nutrient and chlorophyll a data at Futch Creek sampling stations, 1997-1998.
Data as mean (SD) / range, nutrients in mg/L, chlorophyll a as µg/L.
_____________________________________________________________________ Station Nitrate Ammonium Orthophosphate Chlorophyll a _____________________________________________________________________
FC-2 0.005 (0.005) NA 0.006 (0.003) 1.5 (0.9)
0.002-0.018 0.002-0.009 0.4-3.5 FC-4 0.006 (0.004) 0.024 (0.014) 0.006 (0.003) 1.6 (1.0) 0.001-0.012 0.011-0.047 0.002-0.011 0.4-3.3
FC-6 0.009 (0.006) NA 0.006 (0.004) 1.8 (1.3) 0.001-0.020 0.002-0.015 0.4-4.4 FC-8 0.017 (0.011) NA 0.007 (0.005) 1.8 (1.2)
0.005-0.038 0.003-0.021 0.4-4.2
FC-13 0.065 (0.038) NA 0.010 (0.011) 4.1 (3.0) 0.005-0.134 0.002-0.043 0.6-9.2
FC-17 0.136 (0.142 ) 0.051 (0.027) 0.014 (0.020) 6.1 (5.1)
0.018-0.501 0.014-0.083 0.004-0.075 0.8-17.3 FC-20 0.086 (0.079) NA 0.012 (0.020) 5.3 (4.2) 0.033-0.329 0.001-0.073 0.5-13.5
FOY 0.060 (0.055) 0.043 (0.031) 0.009 (0.010) 2.9 (1.5) 0.001-0.209 0.023-0.088 0.003-0.040 0.5-4.8 _____________________________________________________________________
As reportedly previously (Mallin et al. 1996; 1998) the dredging experiment
proved to be successful and the lower portion of the creek was reopened to shellfishing. During 1997-1998 the lower creek maintained excellent microbiological water quality for shellfishing (Table 5.3). The mid-creek areas had good microbial water quality as well, and the uppermost stations continued to have fecal coliform bacterial concentrations
well below those of the pre-dredging period.
Table 5.3. Futch Creek fecal coliform bacteria data, including percent of samples exceeding 43 CFU per 100 mL, 1997-1998. _____________________________________________________________________
Station FC-2 FC-4 FC-6 FC-8 FC-13 FC-17 FC-20 FOY ALL
Geomean 2 2 3 9 15 41 44 11 16 % > 43 /100ml 0 0 0 0 18 55 55 18 18
_____________________________________________________________________
In 1996 we sampled a series of springs and feeder creeks in the upper south branch of Futch Creek. Notable findings (detailed in Mallin et al. 1996) included elevated phosphate concentrations in the feeder branch FC-18C, and nitrate levels as
high as 0.5 mg/L in a spring entering FC-18C. As a follow-up study, in summer 1998
we sampled all of the springs we could locate along the south shore of Futch Creek between FC-8 and FC-25. Orthophosphate and ammonium levels were unremarkable (Table 5.4) but high levels of nitrate (>1.0 mg/L) were found in springs in the upper marsh between FC-13 and FC-17. Two springs in the vicinity of FC-18C again had
nitrate levels near 0.5 mg/L. We emphasize that these springs release nitrate-laden
water into the stretch of Futch Creek between FC-13 and FC-20, where phytoplankton blooms occur in spring and summer (Mallin et al. 1998). We do not know why these particular springs contain elevated nitrate and why others nearby do not (i.e. FC-17S and FC-19s), but it does produce a non-surface source of nitrate to this tidal creek.
Since the N/P ratios were generally high in this area of the creek, phosphorus additions
are likely to cause algal blooms in both spring and summer in this portion of the creek. Table 5.4. Nutrient concentrations (mg/L) in springs along south branch of Futch Creek, June 1998.
_____________________________________________________________________
Station Nitrate Ammonium Orthophosphate _____________________________________________________________________ FC-15S 1.622 0.023 0.013 FC-16S 1.123 0.021 0.032
FC-17S 0.003 0.084 0.013
FC-18S 0.316 0.021 0.011 FC-18HWS 0.421 0.019 0.023 FC-19S 0.001 0.041 0.009 _____________________________________________________________________
6.0 Greenfield Lake One of the major pollution mitigation features in the Greenfield Lake watershed is
an extensive wet detention pond along the Silver Stream branch (Fig. 6.1). The pond
drains approximately 280.5 acres, of which about 43% is impervious surface area. The pond is divided into a 1.25 acre upper and a 1.48 acre lower basin by a causeway pierced by three pipes connecting the flow. In early summer approximately 70% of the upper pond was covered by a mixture of floating and emergent aquatic macrophyte
vegetation, with about 40% of the lower pond covered by vegetation. Principal species
in the upper basin were alligatorweed Alternanthera philoxeroides, pennywort
Hyrocotyle umbellate, water primrose Ludwigia leptocarpa and cattail Typha latifolia, while the lower basin vegetation was dominated by alligatorweed, water primrose and cattail. This pond functioned very well as a nutrient removal system (Table 6.1).
Particularly efficient removal of orthophosphate (83%), nitrate (55%), and ammonium
(54%) was achieved. The decreases in nitrate, ammonium, TN, orthophosphate and TP were all statistically significant. Turbidity and TSS decreased, although this decrease was not statistically significant. Dissolved oxygen significantly increased, probably in part because of aeration while passing through the outfall.
Table 6.1. Comparison of pollutant concentrations in input and output waters of regional wet detention pond on Silver Stream in Greenfield Lake watershed. _____________________________________________________________________
Parameter SS1 SS2
_____________________________________________________________________ DO (mg/L) 5.7 (1.7) 7.5 (1.8)** Turbidity (NTU) 12.0 (23.7) 7.9 (8.7)
TSS (mg/L) 6.3 (12.7) 2.5 (1.8)
Nitrate (mg/L) 0.29 (0.16) 0.13 (0.09)* Ammonium (mg/L) 0.10 (0.06) 0.04 (0.04)* TN (mg/L) 0.84 (0.20) 0.63 (0.18)* Phosphate (mg/L) 0.11 (0.14) 0.02 (0.01)**
TP (mg/L) 0.16 (0.16) 0.04 (0.02)**
Chlorophyll a (µg/L) 5.5 (6.7) 6.5 (5.8) Fecal col. (CFU/100 mL) 11 15 _____________________________________________________________________
* indicates significant difference between input and output concentration at p<0.05
**Indicates significant difference between input and output concentration at p<0.01 Sediment metals were analyzed at one station in Greenfield Lake (GL-2340), one tributary before entering the lake (GL-JRB) and the input (SS1) and output (SS2)
stations of Silver Stream wet detention pond. All sediment metals except mercury were
lower at the pond outfall site than at the input station (Table 6.2). The decreases in aluminum, chromium and iron proved to be statistically significant. While other metals levels were reduced substantially, high variability among the replicate samples for some of the metals (i.e. copper, lead, nickel, zinc) precluded a statistically significant
difference. Regardless, this pond functioned well as a metals removal system.
Table 6.2. Mean and standard deviation of sediment metals concentrations in Greenfield Lake watershed, including a comparison between the input and oufall stations of Silver Stream wet detention pond.
_____________________________________________________________________
Parameter GL-2340 GL-JRB GL-SS1 GL-SS2 _____________________________________________________________________ Al (mg/kg) 854.7 (155.2) 1506.7 (380.8) 704.3 (137.4) 182.7 (38.4)*
As (mg/kg) 0.627 (0.191) 0.242 (0.022) 0.060 (0.048) 0.028 (0.003)
Cd (mg/kg) 0.192 (0.015) 0.184 (0.040) 0.378 (0.257) 0.077 (0.045) Cr (mg/kg) 1.76 (0.18) 2.12 (0.49) 3.68 (0.78) 1.21 (0.23)* Cu (mg/kg) 3.84 (0.34) 5.09 (0.32) 12.58 (7.99) 0.54 (0.41) Fe (mg/kg) 618.7 (896.7) 14.0 (0.67) 211.6 (8.92) 49.2 (1.57)*
Pb (mg/kg) 14.0 (0.67) 12.1 (2.49) 8.92 (3.53) 1.57 (0.42)
Hg (mg/kg) 0.008 (0.006) 0.024 (0.007) 0.010 (0.007) 0.010 (0.006) Ni (mg/kg) 0.77 (0.03) 0.95 (0.21) 1.85 (0.85) 0.30 (0.11) Zn (mg/kg) 16.6 (0.84) 22.3 (2.17) 93.4 (117.6) 1.69 (0.32) _____________________________________________________________________
* Indicates significant difference between input and output concentration at p<0.05
Three tributaries of Greenfield Lake were sampled for physical, chemical, and biological parameters (Table 6.3, Fig. 6.1). Two of the tributaries suffered from extreme hypoxia, with both GL-LB (creek at Lake Branch Drive) and GL-LC (creek beside
Lakeview Commons) both showing hypoxia (DO < 5.0 mg/L) on 10 of 12 occasions.
Turbidity and suspended solids were generally low in the tributary stations (Table 6.3). Nitrate concentrations were high At GL-LC, on several occasions exceeding 1.0 mg/L, moderate at GL-LB and low at GL-JRB (Table 6.3). Ammonium concentrations were highest at GL-LB, exceeding 0.2 mg/L on six of 12 occasions. Orthophosphate
concentrations were high at GL-LC, ranging to moderate at GL-LB. Two of these input
streams maintained fecal coliform levels indicative of poor water quality, with fecal coliform counts exceeding the state standard for human contact waters (200 CFU/100 mL) seven of 12 times at GL-LB and five of 12 times at GL-LC. Chlorophyll a levels were generally low in these streams (Table 6.3).
Three in-lake stations were sampled (Table 6.4). Station GL-2340 represents an area receiving a considerable influx of urban/suburban runoff, GL-YD is downstream and receives some outside impacts, and GL-P is at Greenfield Lake Park, away from inflowing streams but in a high-use waterfowl area (Fig. 6.1). Low dissolved oxygen
was an important factor only at GL-2340, where hypoxia was detected on five of 17
sample trips. Turbidity and suspended solids were low at all three sites. Nitrate concentrations were relatively high at GL-2340, reflecting the proximity of three tributary streams. Nitrate levels decreased considerably toward the park (Table 6.4). Orthophosphate levels were similar at GL-2340 and GL-YD and decreased at the park
(Table 6.4). The changing nitrate dynamics were reflected by the N/P ratios, which
indicated potential limitation of phytoplankton growth by phosphorus at GL-2340, changing to nitrogen limitation at GL-YD and GL-P (Table 6.4). There was also a pattern at all stations of higher N/P ratios in winter and spring and lowest N/P ratios in summer. We have begun a series of nutrient addition bioassay experiments to further
investigate what factors control algal growth in Greenfield Lake. This lake periodically
hosts algal blooms, at times covering large areas of the surface. Our data indicate that blooms may occur during any season, and may consist of either green or blue-green algae, or both at times. Additionally, large blooms of duckweed Lemna sp. often cover
the surface; these blooms are not reflected by chlorophyll a measures as we clear the
surface scum before sampling the subsurface water. Fecal coliform concentrations were generally well within state standards at the in-lake stations (Table 6.4). Table 6.3. Mean and (standard deviation) of water quality parameters in tributary
stations of Greenfield Lake. Fecal coliforms given as geometric mean.
_____________________________________________________________________ Parameter GL-JRB GL-LB GL-LC _____________________________________________________________________ DO (mg/L) 5.3 (1.8) 1.9 (1.6) 3.4 (1.9)
Turbidity (NTU) 6.9 (7.0) 4.8 (3.7) 4.7 (5.5)
TSS (mg/L) 2.5 (1.7) 2.8 (1.8) 3.3 (2.6) Nitrate (mg/L) 0.170 (0.107) 0.283 (0.248) 0.835 (0.225) Ammonium (mg/L) 0.181 (0.309) 0.175 (0.090) 0.101 (0.046) TN (mg/L) 0.992 (0.671) 0.930 (0.247) 1.510 (0.244)
Phosphate (mg/L) 0.057 (0.069) 0.038 (0.016) 0.066 (0.031)
TP (mg/L) 0.090 (0.092) 0.060 (0.023) 0.094 (0.043) N/P molar ratio 12.6 29.9 37.0 Fec. col. (/100 mL) 191 205 166
Chlor. a (µg/L) 3.6 (2.2) 2.0 (1.3) 5.6 (7.9)
_____________________________________________________________________ Table 6.4. Mean and (standard deviation) of water quality parameters in Greenfield Lake sampling stations. Fecal coliforms given as geometric mean.
_____________________________________________________________________
Parameter GL-2340 GL-YD GL-P _____________________________________________________________________ DO (mg/L) 5.9 (2.5) 6.7 (2.0) 8.0 (2.2) Turbidity (NTU) 4.9 (2.7) 3.4 (1.8) 4.6 (7.0)
TSS (mg/L) 4.3 (3.9) 4.1 (2.2) 3.6 (2.0)
Nitrate (mg/L) 0.196 (0.170) 0.075 (0.077) 0.013 (0.012) Ammonium (mg/L) 0.055 (0.049) 0.034 (0.039) 0.028 (0.047) TN (mg/L) 0.980 (0.340) 0.841 (0.230) 0.625 (0.206) Phosphate (mg/L) 0.018 (0.013) 0.020 (0.015) 0.012 (0.008)
TP (mg/L) 0.045 (0.024) 0.048 (0.021) 0.036 (0.018)
N/P molar ratio 24.9 11.4 4.6 Fec. col. (/100 mL) 93 21 76
Chlor. a (µg/L) 11.7 (13.2) 17.4 (12.7) 20.5 (24.9)
____________________________________________________________________
7.0 Hewletts Creek Hewletts Creek was sampled at four tidally-influenced areas (FC-2, NB-GLR,
MB-PGR and SB-PGR) and one freshwater runoff collection area near Longleaf Mall
(HC-LO - Fig. 7.1). In addition to the standard sampling described in the Methods, nutrient addition bioassays were performed on water from HC-2 and SB-PGR during spring and summer. These bioassays consisted of collecting creek water in 20-L carboys, transporting them to the laboratory, and dispensing 3 L into each of 30 gallon-
sized cubitainers. The cubitainers were then spiked with different nutrients (nitrate-N as
100 ppb (µg/L), nitrate-N as 50 ppb, phosphate-P as 50 ppb, phosphate-P as 25 ppb, and no additions, which was the control). All treatments were in triplicate. The cubitainers were incubated outdoors in pools covered with 2 layers of neutral-density
screening to avoid photostress. The water and cubitainers were kept gently agitated
with an aquarium pump. Chlorophyll a samples were collected daily for three days, with results compared statistically using SAS. Treatments yielding significantly higher chlorophyll a (p < 0.05) were considered to contain the nutrient most limiting to phytoplankton growth.
Physical data indicated that turbidity on average exceeded the state standard for estuarine water at both NB-GLR and SB-PGR (Table 7.1), although turbidity was only assessed on four occasions. Dissolved oxygen remained above standard at all creek stations except for hypoxic conditions in July at NB-GLR and SB-PGR. Nitrate
concentrations were high in the middle branch (MB-PGR) which drains both Pine Valley
and the Wilmington Municipal Golf Courses (Fig. 7.1; Mallin and Wheeler 1998). Both NB-GLR and SB-PGR also periodically receive elevated nutrient loading as well. Nitrate, ammonium and orthophosphate data were used to compute inorganic nitrogen to phosphorus molar ratios (Table 7.1). Median N/P rations demonstrated that both HC-
2 and SB-PGR should be primarily nitrogen limited, while NB-GLR tends toward
potential phosphorus limitation. Our spring and summer experiments confirmed that phytoplankton growth at HC-2 was limited by nitrogen, while phytoplankton growth at SB-PGR was limited by nitrogen in summer (Fig. 7.2). In spring algal growth at this station was not nutrient-limited; rather, it was probably light-limited (Fig. 7.2).
Chlorophyll a data (Table 7.1) showed that Hewletts Creek continued to host severe algal blooms, as it has in the past (Mallin et al. 1998a). During the past year the state chlorophyll a standard of 40 ppb was exceeded in June at NB-GLR and in February, June and July at SB-PGR. These phytoplankton blooms were probably the
principal cause of the elevated turbidity levels at SB-PGR. From our experimental data
and the computed N/P ratios, the blooms at NB-GLR and SB-PGR are probably caused by periodic pulses of nitrate entering the north and south branches of Hewletts Creek. The most likely source of these inputs is non-point source runoff from lawns, gardens, and land-disturbing activities. While the highest in-stream nutrient levels are at MB-
PGR, the heavily shaded nature of the stream at this station precludes formation of
phytoplankton blooms. However, nutrients carried downstream reach open waters, including the area where the middle and north branches join.
Table 7.1. Selected water quality parameters at tidally-influenced stations in Hewletts Creek watershed as mean (standard deviation) / range. _____________________________________________________________________
Parameter HC-2 NB-GLR MB-PGR SB-PGR
_____________________________________________________________________ Salinity 28.0 (8.8) 9.0 (12.4) 1.2 (2.7) 13.9 (12.2) 6.7-37.0 0.1-33.4 0.0-6.7 0.3-35.2
Turbidity (NTU) 19.7 (10.8) 26.2 (9.4) 11.3 (14.7) 26.4 (5.6) 7.6-32.1 17.2-36.5 1.3-28.1 20.6-32.6 DO (mg/L) 7.5 (1.4) 7.5 (1.9) 6.7 (1.1) 7.2 (2.0)
5.6-9.2 4.5-10.0 5.4-7.8 4.2-11.0
Nitrate (mg/L) 0.013 (0.018) 0.100 (0.067) 0.232 (0.089) 0.039 (0.033) 0.001-0.062 0.005-0.215 0.070-0.390 0.002-0.098
Ammonium (mg/L) 0.027 (0.020) 0.039 (0.046) NA 0.014 (0.013)
0.008-0.065 0.001-0.105 0.001-0.033 Phosphate (mg/L) 0.008 (0.007) 0.013 (0.010) 0.014 (0.010) 0.012 (0.009) 0.002-0.023 0.002-0.036 0.005-0.038 0.001-0.032
Mean N/P ratio 13.3 (12.5) 38.1 (37.4) NA 37.3 (71.9) Median 9.5 22.9 6.3 Chlor a (ug/L) 1.6 (1.1) 12.7 (23.7) 4.1 (7.8) 29.2 (45.8)
0.4-4.0 1.0-87.0 0.2-27.4 0.7-126.7
_____________________________________________________________________
Data from the freshwater collection station at HC-LO near Longleaf Mall did not show any remarkable pollutant levels during the sampling period of October 1997-July 1998 (Table 7.2). Water from this location drains into the middle branch of Hewletts
Creek; however, nutrient levels are considerable lower than those at MB-PGR (Table
7.1) indicating that HC-LO was not an important nutrient source to Hewletts Creek under normal circumstances. Table 7.2. Parameter concentrations at freshwater collection station near Longleaf Mall
as mean (SD) and range.
_____________________________________________________________________ Parameter HC-LO _____________________________________________________________________
DO (mg/L) 5.9 (1.2) 3.6-7.7 Turbidity (NTU) 8.8 (8.9) 2.6-31.5 TSS (mg/L) 2.6 (1.7) 0.5-6.0 Nitrate (mg/L) 0.067 (0.026) 0.020-0.110
Ammonium (mg/L) 0.126 (0.040) 0.070-0.200
TN (mg/L) 0.680 (0.133) 0.500-0.870 Phosphate (mg/L) 0.018 (0.011) 0.005-0.030 TP (mg/L) 0.035 (0.011) 0.020-0.050
Chlorophyll a (µg/L) 2.2 (2.2) 0.6-7.7
Fecal col. (CFU/100 mL) 35 3-112 _____________________________________________________________________
8.0 Howe Creek Water Quality Howe Creek was sampled for physical parameters, nutrients, and chlorophyll a at
three locations during 1997-1998 (HW-FP, HW-GC, and HW-GP, Fig. 8.1). Howe
Creek, as in past years, portrayed a tidal creek with good water quality near the ICW and poor water quality in the oligohaline to mesohaline regions. Turbidity is low near the ICW but exceeded State standards on several occasions in the upper reaches (Table 8.1). The upper Howe Creek watershed maintains several land clearing and
construction projects which appear to affect the water quality by non-point source
sediment runoff. All three of the stations had incidences of substandard (< 5.0 mg/L) dissolved oxygen (Table 8.1), with HW-GP having the most incidents (four). As mentioned, both turbidity and low dissolved oxygen impair the fisheries primary nursery function of an estuary. As development continues in the upper watershed, we expect
continuing degradation of upper and middle Howe Creek from non-point source runoff.
Nutrient levels are low near the ICW but can be elevated in the creek near Graham Pond (Table 8.1). Algal bloom conditions occur periodically in spring and summer at Station HW-GP (Table 8.1; Mallin et al. 1998). Median inorganic molar N/P
ratios are low, indicating that nitrogen is probably the principal limiting nutrient at HW-
FP and often at HW-GP, although periodic nitrate loading events will drive the ratio upward at times in the upper stations (Table 8.1). One nutrient source was uncovered during the sample year which may contribute to algal bloom formation at times. There is a blowdown pipe from the municipal water supply which discharges into a runoff ditch
(LF-RO) which enters the creek near HW-GP (Fig. 8.1). Nitrate and phosphate levels in
this water are high in ecological terms (0.80 mg/L and 0.36 mg/L, respectively), due to source water and chemicals added in the water treatment process (D. Bradshaw, pers. comm.). It is our recommendation that this blowdown be moved to an area that discharges into the wetland in upper Graham Pond so nutrients will be removed by
biological processes before reaching the more sensitive estuarine waters.
Table 8.1. Selected water quality parameters in Howe Creek as mean (standard deviation) / range. _____________________________________________________________________
Parameter HW-FP HW-GC HW-GP
_____________________________________________________________________ Salinity 29.8 (8.7) 25.9 (8.4) 13.9 (12.5) 7.2-37.0 8.7-35.2 0.2-33.6
Turbidity (NTU) 6.8 (2.6) 15.9 (7.6) 31.9 (34.9) 2.9-10.6 2.6-25.7 3.2-100.0 DO (mg/L) 6.6 (2.0) 6.6 (2.0) 6.5 (2.3)
3.6-9.2 3.0-9.2 2.9-9.4
Nitrate (mg/L) 0.006 (0.007) 0.012 (0.017) 0.038 (0.050) 0.001-0.023) 0.001-0.055 0.001-0.162
Ammonium (mg/L) 0.019 (0.013) NA 0.013 (0.012)
0.001-0.037 0.001-0.029 Phosphate (mg/L) 0.006 (0.002) 0.008 (0.009) 0.008 (0.005) 0.003-0.009 0.002-0.035 0.001-0.019
Mean N/P ratio 9.7 (7.6) NA 32.3 (55.4) Median 7.0 8.1
Chlor a (µg/L) 2.8 (3.1) 2.4 (1.3) 12.5 (17.8)
0.5-11.8 0.5-4.7 0.5-56.3 _____________________________________________________________________
8.1
Tidal Flushing of Howe Creek
Jason Hales and Lawrence B. Cahoon Department of Biological Sciences UNC Wilmington
Introduction The success of the 1996 dredging project at Futch Creek, in which two channels were dredged through a sand bar that had built up across the mouth of the creek along the
IntraCoastal WaterWay, has opened the possibility that this approach may be useful in
improving water quality in other tidal creeks. Water quality studies in Futch Creek had previously established that levels of fecal coliform bacteria were high enough to force shellfishing closures throughout the creek, but that no manageable sources of fecal coliform bacteria could be found. Therefore, an effort to improve flushing and dilution of the
creek's waters by dredging its mouth was approved and executed. The success of this
approach was evidenced by improvements in fecal coliform levels sufficient to allow reopening of shellfishing in the lower portion of the creek (Mallin et al., 1996; 1998). The situation in Futch Creek (restricted flushing by outside water resulting in elevated fecal coliform concentrations) that led to a decision to try dredging may occur in other tidal
creeks as well, suggesting that limited dredging may have wider application as a water
quality management tool. Furthermore, the problems arising from potential closure of inlets, especially Mason's Inlet, suggest that flushing and dilution of pollutants are considerations in that context as well. One of the factors critical to a determination that dredging may improve water
quality is an analysis of flushing rates of a tidal creek under current conditions. If water
quality is poor but flushing rates are high, then management of pollution sources is likely to be more cost-effective than dredging for restoring water quality. Conversely, if flushing rates are very low, then only a very extensive (and therefore expensive and difficult to permit) dredging project might increase flushing rates sufficiently to improve water quality
in a meaningful way.
We measured flushing rates in Howe Creek, a tidal creek that has been closed to shellfishing since the early 1990's owing to elevated fecal coliform counts. Our aim was to determine flushing rates and to compare those rates to values found in other similar studies of Futch Creek and Bald Head Creek.
Methods Flushing rates were calculated by measuring the concentrations of rhodamine dye over time at several locations in Howe Creek (Fig. 1), following a period of dye
introduction. Rhodamine dye is commonly used for this kind of measurement, as it is very soluble, non-toxic, and brightly fluorescent, allowing its quantification to part per billion concentrations using a properly configured fluorometer.
Rhodamine dye (60 x 106 ppB) was dispensed into Howe Creek at Site D (Fig. 1) at
a rate of 145 ml/min for 110 min during a mid-ebbing tide on July 13, 1998. Triplicate 50 ml water samples were then collected at each site (A-D) at ebb tide for the next eleven tidal cycles. Samples were analyzed by fluorometry using a Turner 450 fluorometer with local water collected before dye introduction used as blanks and known concentrations of dye in
Howe Creek water as standards.
Dilution rates and dilution times were calculated using dye concentrations over time for each site and an exponential decay equation of the form Ct = Coe-kt
where Ct is the dye concentration at time t, Co is initial dye concentration, e is the base of the natural log, k is the dilution rate constant and t is the time interval. Values of k and t are then determined by rearranging the equation using known values of dye concentration at different times.
Concurrent measurements of salinity were made using a refractometer, and
observations of flow and tidal heights were made to determine actual ebb tide at each site. Results Fig. 2 illustrates results of measurements of dye concentration and salinity at site B
(chosen as an example of the data generated from each site) in Howe Creek.
A rain event (1.5" reported at Wrightsville Beach) during the middle of the dye study resulted in a notable decline in salinity (Fig. 2) and allowed estimation of dilution rates in Howe Creek under both dry and runoff event conditions. Flushing rates in Howe Creek before the rain event ranged from 44% per tidal cycle at sites B and D to 48% at site C and
49% at site A. After the rain event, which drove noticeable amounts of stormwater runoff
into Howe Creek, flushing rates declined to 30% at site A, 25% at site B, 15% at site C and 24% at site D. Discussion
Flushing rates in Howe Creek during no-rain conditions were relatively high (range 44-49% per tidal cycle) and suggest that Howe Creek can be effectively flushed. In comparison, measures of flushing rates in Futch Creek after the 1996 dredging project were in the range of 40-47% per tidal cycle (Hales, 1998). As described above, flushing
rates in Futch Creek were high enough after dredging to permit reopening to shellfishing. This comparison suggests that Howe Creek, which is currently closed to shellfishing, is nearly well enough flushed to benefit from a dredging project. The significantly lower flushing rates measured in Howe Creek after a rain event (15-30% per tidal cycle) indicate that a strong pulse of freshwater runoff changes the
fundamental properties of this small tidal estuary in a way that favors retention of stormwater runoff and the pollutants it carries. It is likely that strong pulses of freshwater create a stratified condition that inhibits mixing with the saltier water normally resident in Howe Creek. Formation of a freshwater lens in the creek itself may also prevent effective flushing by forming a front at the juncture with saltier IntraCoastal Waterway water. In any
event, the reduced flushing rates associated with a rain event indicate that dredging alone may be insufficient to improve water quality reliably enough to permit reopening the creek to shellfishing. Stronger efforts to manage stormwater runoff from the rapidly developing Howe Creek drainage basin will likely have to be made before significant water quality improvements can be attained in the creek.
9.0 Pages Creek Pages Creek was sampled at three stations, of which two receive drainage from
developed areas (PC-BDUS and PC-BDDS) and a well-flushed one near the ICW (PC-
M - Fig. 9.1). This is one of the least-polluted tidal creeks in New Hanover County (Mallin et al. 1998). During the past sample year turbidity was low to moderate with only one incident of turbidity exceeding the state standard of 25 NTU (Table 9.1). There were several incidents of hypoxia during summers of 1997 and 1998, primarily in the
stations draining Bayshore Drive. Nutrient concentrations were normally low, and
phytoplankton biomass was low to moderate (Table 9.1). Inorganic nitrogen-to- phosphorus molar ratios were below 16, indicating that phytoplankton growth in this creek is probably nitrogen limited.
Table 9.1. Selected water quality parameters in Pages Creek as mean (standard
deviation) / range. _____________________________________________________________________ Parameter PC-M PC-BDDS PC-BDUS _____________________________________________________________________
Salinity 32.1 (3.8) 21.2 (13.8) 16.8 (9.3) 23.0-36.5 1.6-35.7 2.2-29.4 Turbidity (NTU) 7.6 (3.0) 14.1 (9.4) 12.1 (7.5)
4.9-12.3 2.1-27.5 2.3-21.4
DO (mg/L) 7.1 (1.8) 7.1 (2.7) 6.3 (2.3) 4.2-9.9 2.3-10.6 1.7-9.7
Nitrate (mg/L) 0.005 (0.038) 0.031(0.04) 0.017(0.015)
0.001-0.014) 0.029-0.124 0.001-0.047 Ammonium (mg/L) 0.012 (0.009) 0.032 (0.013) 0.054 (0.043) 0.001-0.025 0.017-0.058 0.032-0.128
Phosphate (mg/L) 0.006 (0.002) 0.007 (0.006 ) 0.012 (0.013) 0.004-0.009 0.002-0.022 0.002-0.047 Mean N/P Ratio 5.9 (4.1) 30.6 (45.2) 15.3 (12.9)
median 7.1 12.1 14.4
Chlor a (ug/L) 1.9 (1.4) 3.6 (3.2) 6.8 (7.9) 0.4-4.4 0.5-10.0 0.4-26.5 _____________________________________________________________________
10.0 Smith Creek The three stations sampled in this watershed were SC-GT, a tributary of Smith
Creek draining a highway construction area (Smith Creek Parkway), and two estuarine
sites on Smith Creek proper, SC-23 and SC-CH (Fig. 10.1). The North Carolina turbidity standard for estuarine waters (25 NTU) was exceeded on a number of occasions at all stations, and the two Smith Creek stations had overall mean turbidities exceeding the standard (Table 10.1). Total suspended solids were periodically high at SC-GT,
probably as a result of construction activities. On average, most nutrient concentrations
were unremarkable, and there was a tendency for concentrations to increase downstream toward the Cape Fear River (Table 10.1). Fecal coliform bacteria levels exceeded the North Carolina standard for human contact waters (200 CFU/100 mL) at times at all stations. The geometric mean fecal coliform concentration was below the
human contact standard at all three stations but well above the shellfishing standard (14
CFU/100 mL) in the estuarine portion of the creek (Table 10.1). Table 10.1. Selected water quality parameters in Smith Creek watershed as mean (standard deviation) / range.
_____________________________________________________________________
Parameter SC-GT SC-23 SC-CH _____________________________________________________________________ Turbidity (NTU) 27.4 (27.8) 36.8 (30.4) 27.1 (18.2)
2.3-94.0 10.0-100.9 8.5-62.2
TSS (mg/L) 21.6 (37.3) 11.3 (3.2) 13.2 (8.0) 2.5-124.0 7.0-16.0 3.0-28.0
Nitrate (mg/L) 0.04 (0.03) 0.19 (0.12) 0.22 (0.09)
0.01-0.05 0.10-0.51) 0.07-0.36 Ammonium (mg/L) 0.04 (0.02) 0.05 (0.03) 0.06 (0.04) 0.01-0.06) 0.02-0.13 0.01-0.14
Phosphate (mg/L) 0.04 (0.04) 0.06 (0.03) 0.06 (0.02) 0.01-0.12 0.02-0.12 0.03-0.10 Fecal col. /100 mL 90 58 112
(geomean / range) 1-800 17-235 6-485
_____________________________________________________________________
11.0 Upper and Lower Cape Fear Within the Wilmington City limits drainage directly to the Cape Fear River (CFR)
was sampled at one location each in the Upper and Lower Cape Fear Watersheds. The
stream draining the Upper CFR was sampled behind the Wilmington Police Station between 2nd and 3rd Sts (Fig. 11.1). Concentrations of most physical, chemical and biological constituents were low to moderate except for nitrate (Table 11.1). Nitrate concentrations in this stream were the highest in any of the City or County drainage’s.
The source of this nitrate is currently not defined, but we plan to conduct further
investigations of possible sources. Drainage from the Lower CFR was sampled from the stream draining Greenfield Lake (Fig. 11.2). Processing within the lake served to keep concentrations of most
constituents relatively low (Table 11.1). Average fecal coliforms, turbidity, and
chlorophyll a were all well below state water quality standards during the sampling period. Table 11.1 Water quality summary statistics (mean (standard deviation) / range) for
Wilmington Upper (UCF) and Lower (LCF) Cape Fear Watersheds.
_____________________________________________________________________ Station DO (mg/L) Turbidity (NTU) TSS (mg/L) Fecal col (CFU/100 mL) _____________________________________________________________________
UCF 8.2 (0.4) 0.6 (0.7) 1.4 (0.9) 69 7.3-8.8 0.0-2.3 0.5-3.0 21-300 LCF 8.4 (2.2) 2.0 (1.5) 2.4 (2.1) 24
4.0-10.9 0.0-3.9 0.5-7.0 3-105
Nitrate (mg/L) Ammonium (mg/L) Orthophosphate (mg/L) _____________________________________________________________________
UCF 3.960 (2.512) 0.015 (0.011) 0.029 (0.011)
2.500-11.000 0.005-0.040 0.010-0.050 LCF 0.028 (0.045) 0.027 (0.019) 0.014 (0.009) 0.005-0.150 0.005-0.050 0.050-0.030
_____________________________________________________________________
Sediment metals were sampled at UCF-PS (Table 11.2). Most of the concentrations were similar to those found elsewhere in the system, but several metals had unusually high concentrations at this location. Lead and arsenic levels were
highest in the system here, and iron was elevated as well (Table 11.2). Lead was at the
approximate concentration in which harmful effects to aquatic life will occasionally occur, and arsenic concentrations were close to that category (Appendix A).
Table 11.2. Mean and standard deviation of sediment metals concentrations at
Station UCF-PS in the Upper Cape Fear watershed.
_____________________________________________________________________ Parameter (mg/kg) UCF-PS _____________________________________________________________________ Al 1192.67 (588.87)
As 6.303 (3.846)
Cd 0.245 (0.134) Cr 1.48 (0.68) Cu 8.08 (4.51) Fe 2276.7 (1047.4)
Pb 46.2 (43.9)
Hg 0.006 (0.002) Ni 1.16 (0.53) Zn 31.83 (24.01) _____________________________________________________________________
12.0 Pfiesteria Analysis in the Tidal Creeks In 1997 a series of phytoplankton samples was taken in an effort to assess the
distribution of the potentially-toxic dinoflagellate Pfiesteria piscicida in the tidal creek
system. Previous samples showed it was present in these creeks (Mallin et al. 1996; 1998) although no Pfiesteria-induced fish kills have been reported. Monthly from May through September we sampled a series of 13 stations that represented extremes in water quality, at or near high tide. Samples for Pfiesteria were collected in 60 mL dark
bottles and preserved with Lugol’s iodine solution, and sent to the Aquatic Botany
Laboratory at North Carolina State University for identification. Counts were done using Palmer-Maloney cells. Concurrent samples for nitrate, ammonium, total nitrogen, orthophosphate, total phosphorus, chlorophyll a, water temperature, salinity, dissolved oxygen and turbidity were also collected with the Pfiesteria samples. The completed
data were logged into SAS and correlation analyses were performed to assess any
environmental variables that may prove important in explaining non-toxic Pfiesteria distribution. Pfiesteria cells were seen in 12% of the samples taken. No Pfiesteria were
found in either May or September. During June, July and August Pfiesteria cells were
found in several locations, including the north and south branches of Bradley Creek at Wrightsville Avenue (BC-NB and BC-SB – Fig. 3.1), the south branch of Futch Creek (FC-17 – Fig. 5.1), the south branch of Hewletts Creek at Pine Grove Road (SB-PGR – Fig. 7.1), and both stations abutting Bayshore Drive in Pages Creek (PC-BDDS and PC-
BDUS – Fig. 9.1) It is notable that these stations show periodic signs of eutrophication
including elevated nutrient loading and periodic algal blooms (Mallin et al. 1998). The August Pfiesteria ‘hits’ only occurred where there were elevated numbers of cryptomonads, flagellated phytoplankton cells which serve as a food source for Pfiesteria (Burkholder and Glasgow 1997).
Pfiesteria counts were highly significantly correlated with chlorophyll a and inversely with salinity. Counts were also significantly correlated with total phosphorus and nitrate concentrations (Table 12.1). Since Pfiesteria utilize phytoplankton as a food source, the correlation with chlorophyll a is not surprising. Pfiesteria can utilize organic
phosphorus directly, and a large percentage of the total phosphorus in these creeks is
organic. Nitrate stimulates phytoplankton growth in these creeks; thus, it indirectly stimulates Pfiesteria growth. The inverse relationship with salinity is not to imply that
Pfiesteria only thrive in low salinity waters; rather, most of the highest salinity stations in this creek complex maintain the lowest nutrient and chlorophyll a levels.
The creek stations that host Pfiesteria tend to be those which display the classic habitat noted in Burkholder et al. (1995), of shallow, poorly-flushed estuarine areas receiving nutrient inputs. These collections were done at or near high tide, and we demonstrated previously that both higher chlorophyll a and higher Pfiesteria counts
occur at or near low tide (Mallin et al. 1996). Therefore, the results of this survey should
be considered to be conservative. We emphasize that the Pfiesteria found were in non-toxic stages, and we are unaware of any fish kills in these creeks that have been attributed to Pfiesteria. However, we also emphasize that since certain areas of these creeks do support Pfiesteria, at times in elevated concentrations (Mallin et al. 1996),
there remains a potential for future Pfiesteria-induced fish kills.
Table 12.1. Results of correlation analyses between Pfiesteria counts and physical,
chemical, and biological parameters in New Hanover County tidal Creeks, summer
1997. Pfiesteria, chlorophyll a, and nutrient data were normalized by log-transformation. Pearson correlation coefficient (r) / probability (p).
SAL CHLA TP NIT
PP -0.495 0.497 0.284 0.288 0.001 0.001 0.022 0.020
PP = Pfiesteria counts, CHLA = chlorophyll a concentration, TP = total phosphorus
concentration, NIT = nitrate concentration
13.0 Tidal Creeks Benthic Community Clams, polychaete worms, amphipods, crabs and other crustaceans are some of
the major groups that make up the benthic community in most estuarine systems.
These organisms are important links in the food chain in that they crop primary production and transfer that energy to higher other trophic groups. Assessments of benthic communities are included in a number of environmental monitoring programs, such as EMAP (EPA’s Environmental Monitoring and Assessment Program) and long
term studies looking at the health of specific estuaries (Chesapeake Bay Program,
Lower Cape Fear River Program, and Florida Bay project, among others). Benthic studies provide information about the workings of the estuarine system, that when combined with water quality data (instantaneous measures for a number of parameters), give insight into the duration and fate of pollutants and nutrients within the
system. Among the benthic fauna, some groups have developed life history strategies
to take advantage of disturbances. These organisms tend to do best in areas subjected to regular disturbances and may be out competed in other areas. Likewise there are other groups that tend to do better in areas with less disturbance and better water quality conditions. These organisms tend to be slower growing and more susceptible to
environmental changes. It is the presence and proportions of these different types of
taxa that allow us to evaluate the condition and stability of the estuarine environment. The Cape Fear River Estuary is a good example of this point. Following Hurricane Fran, the benthic community was greatly impacted, with decreased abundance at all sites. The monthly monitoring following the hurricane showed increased abundance of
a few taxa (those tolerant of low oxygen levels and high siltation). Over the next year
species richness rose and the benthic community returned to near pre-storm levels. These types of community indicators are a good benchmark of the average conditions experienced at a particular site and have the added advantage of allowing some conclusions about environmental conditions independent of daily water quality
measurements.
As part of the New Hanover County Creeks Project the Benthic Ecology Lab has been looking at the benthic biota of the tidal creeks within the county since 1993. Assessment of the bivalve community within these creeks has been a part of this
ongoing project. The bivalve community was chosen because of its close relationship
with sediment quality and limited mobility in general (although there are some species of bivalve capable of large movements). Juvenile bivalves are particularly susceptible to changes in sediment and nutrient loading. Increases in either of these factors could lead to interference with feeding or burial causing mortality. By monitoring the
composition and abundance of the bivalve community we can see the effects of
activities within the creek systems. With the existing data set, we can now separate indicators of more serious environmental change, in particular changes in species richness and/or abundance or changes in community composition, from inter-annual variability that is part of the natural cycle.
In addition to bivalve surveys, habitat utilization by fish and crabs within Hewletts and Pages Creeks was examined in the summer 1997 as part of a related but independent project. This study examined differences in habitat utilization with increasing habitat complexity. The composition and abundance of fish and crabs
utilizing various marsh habitats within these particular mini-estuaries is critical to
understanding the health and function of these systems. Habitat comparisons were made among solitary marsh grass, oyster, and contiguous marsh grass and oyster stands. Hewletts and Pages Creeks represent very different water quality situations
within New Hanover County. Pages Creek is considered to have some of the highest
water quality, while Hewletts Creek is considered to be one of the most impacted of the tidal creeks within the county (Mallin et al. 1994; 1998). A bivalve growth study was initiated in the 1997-98 sampling season at Hewletts
and Pages Creeks. Because Hurricane Bonnie destroyed part of the experiment,
results are inconclusive. Bivalve Surveys
Bivalve surveys were conducted in the fall of 1997, following the same general
sites and methods as in previous years. Five replicate 1 m2 plots were chosen haphazardly from a research site at the mouth of Hewletts, Bradley, Howe, and Pages Creeks. The sediment characteristics among sites are very similar, being dominated by fine sands. These sites were representative of tidal flats in the lower estuary, and of
tidal flats in marsh systems throughout southeastern North Carolina, and all sites were within 100 m of the Intracoastal Waterway. Each plot was excavated to a depth of 20 cm and the contents passed through a 2 mm screen. All bivalves retained were identified and measured (length and width).
Twelve bivalve taxa were recovered from excavations in the fall 1997 surveys. This is comparable with previous sampling seasons. Tagelus plebius was the most common taxa recorded, with highest numbers found in Howe Creek (Table 13.1). There was no difference in the total abundance of bivalves recovered from each creek (Fig. 13.1) and diversity measures reveal Howe has the lowest diversity (0.09) and Hewletts
the highest (Fig. 13.2). Low diversity at Howe Creek reflects the fact that only four taxa were found there, Tagelus dominating (Table 13.1). The other three creeks had more overall taxa recovered (9-10), with a more even distribution. These findings follow patterns seen in previous years (Mallin et al. 1998), although it should be noted that abundance patterns are not consistent among all years. Size measurements are
difficult to interpret because of the variability in the number of individuals per species (Table 13.2). Habitat Utilization
The following sections are related to habitat utilization, both Breder trap and pit trap studies are independent of the creeks project (Rhoads 1998). However the data is related to community composition within the creeks and is useful in considering habitat function.
Breder Traps In order to test habitat utilization by epibenthic organisms, Breder traps and pit
traps were set in intertidal areas of solitary marsh cordgrass, oyster, and contiguous
marsh grass and oyster habitats in Hewletts and Pages Creeks. Breder traps are clear acrylic boxes (31 cm X 15 cm X 16 cm) with a 15 cm long funnel at one end. There is a 2.5 cm wide opening in the funnel to allow entrance of organisms. Three traps were placed at least 0.5 m into each habitat and allowed to fish for 2 hours (traps were set
three times during the summer of 1997 from June to Sept.). One of each of the four
habitat types was sampled simultaneously on a given day. A total of 15 replicates of each habitat type were sampled, for a total of 60 patches in each creek. This sampling method targets small near-bottom nekton (Sargent and Carlson, 1987; Webb, 1987; Townsend, 1991) that are considered transient users of the habitat. All fish caught in
these traps were identified and measured (total length).
Pinfish (Lagodon rhomboides) and the mummichog (Fundulus heteroclitus) comprised 90% of all fish caught in Breder traps. Pinfish show a shift in habitat utilization between creeks with isolated oyster used more heavily in Hewletts Creek and
isolated oyster and isolated Spartina used in Pages Creek at the beginning to the
summer. However pinfish switched to isolated Spartina patches in both creeks by the end of the summer (Rhoads 1998). This could reflect shifts in food resources and possibly growth of pinfish throughout the sampling period.
Mummichogs showed a different utilization pattern with larger individuals caught
in the beginning of the summer. Hewletts Creek showed no significant difference among the four habitat types in the first sampling period switching to mixed oyster in the second sampling and mixed Spartina in the third. Mixed Spartina and isolated oyster were more heavily used at Pages Creek in the beginning switching to isolated Spartina
and then to mixed and isolated Spartina. It should be noted here that mummichogs
were absent from the creeks for several weeks in the middle sampling period and that a smaller size class returned to the creeks at the end of the summer. F. heteroclitus utilized all habitats sampled, although there was variation through time and space (creeks). There was a clumped distribution, probably reflecting schooling behavior.
What we do not know is if these habitats are complimentary or if use is independent (i.e.
multiple habitats are utilized for particular resources and/or refuge or they are interchangeable for a specific function). Pit Traps
Pit traps are plastic containers (38 cm X 26 cm X 15 cm) that were buried flush with the sediment surface. This sampling was conducted twice during the summer of 1997. A single pit trap was placed 1 m within each habitat patch (15 replicate patches for each habitat type, Spartina, oyster, mixed Spartina, and mixed oyster) and left for
two weeks. At the end of this time the traps were retrieved and the contents sieved
through a 5 mm screen. In order to simulate the appropriate habitat, each container was filled with ambient sediment and either oyster or artificial marsh grass (6 mm diameter dowel rod to simulate vegetation stems). This sampling method targets resident shrimp and crab fauna. All crabs and shrimp retained were identified and
measured.
The most common species caught in pit traps were two species of mud crab, Panopeus herbstii, Rhithropanopeus harrissi, and several species of fiddler crab (Uca
spp.). Panopeus herbstii was more common in isolated and mixed oyster over isolated
or mixed Spartina in both creeks. Interactions occurred among creek, habitat and time period. Rhithropanopeous harrissi showed strong habitat preferences with significantly greater abundance in mixed and isolated oyster compared to mixed Spartina, which in turn was significantly greater than isolated Spartina. Uca spp., were more abundant in
Hewletts Creek compared to Pages Creek. Isolated oyster was used significantly more
by Uca than isolated Spartina or mixed Spartina, which were used more than mixed oyster at Hewletts Creek. This pattern changes slightly at Pages Creek, with mixed
Spartina and isolated Spartina used more than mixed oyster or isolated oyster. Overall mud crabs (P. herbstii and R. harrissi) utilize oyster to a greater extent, while Uca use
all habitats sampled to some degree.
Growth Experiment Because differences in water quality could affect growth and survival of juvenile
bivalves, Hewletts and Pages creeks were chosen as locations for study of growth
under differing nutrient and productivity regimes. Small Mercenaria mercenaria (obtained from Sea Perfect Hatchery) were measured and placed into grow-out bags. Four hundred clams were placed in each grow-out bag. Two grow-out bags were set in each creek (attached to the bottom with stakes). Clams were placed in August of 1998
and allowed to grow for 6 weeks before retrieval. After retrieval all surviving clams were
measured again. Although this experiment was interrupted by Hurricane Bonnie preliminary data showed mean growth up to 3.41 mm in five weeks. This experiment will be repeated in the spring and fall of 1999.
Discussion
There are a number of factors working within the tidal creeks of New Hanover County that could impact bivalve distributions as well as how epibenthic organisms utilize these habitats. It is doubtful that larval supply would explain the difference seen
among the tidal creeks because all creeks are near equal distance from inlets and flow
characteristics do not correlate with abundance patterns. Changes in development along the creeks is potentially a significant factor, since increases in runoff (nutrient as well as sediment) can affect juvenile bivalves through interference with feeding and suffocation by burial. In previous years, samples were also collected in the upper
regions of each creek but no live clams were ever recovered. Changes in sediment
characteristics, possibly caused by increased development and runoff, was one suggested explanation. It is important to continue to follow the trend seen in the lower areas of each creek, because of the potential for similar impacts in those areas. Our previous work has allowed us to evaluate inter-annual variations in these mini-estuarine
systems. Through continued monitoring we can detect changes in the existing patterns
of distribution and abundance.
14.0 Use of Semi-permeable Membrane Devices (SPMDs) For the Detection of Organic
Pesticide, Herbicide, and Fungicide Residues in Howe Creek,
Wilmington, North Carolina Dr. John J. Manock, Richard C. Schibetta Department of Chemistry
University of North Carolina at Wilmington
Two SPMD devices were deployed at the mouth of Howe Creek near the Jack Nicklaus golf course in Wilmington, NC, for a period of 27 days. The average temperature during the exposure period was 28.00C. SPMDs were processed at
Environmental Sampling Technologies laboratory in Missouri and extracts subjected to
microcolumn clean up with silica. Analysis by GC and GC/MS found many fatty acids and hydrocarbons, but no trace of organic pesticides. In addition, the CMLS program simulated the movement of five chemicals applied to the course during the exposure period to a depth of one meter. Each report consisted of 250 simulations in specific
horizons of Seagate, Murville, and Baymeade Fine Sands for glyphosate, fenoxycarb,
diquat dibromide, bifenthrin, and chlorpyrifos. The amount of chemical reaching a depth of one meter was also determined. Of the five chemicals, glyphosate, fenoxycarb, and diquat dibromide were found to have similar behavior in the three soils in respect to both travel times and amounts reaching one meter. Chlorpyrifos and bifenthrin also showed
similar patterns in both areas of the model. In general, the simulations suggest that
glyphosate, fenoxycarb, and diquat dibromide have the best chance to move quickly through the sediments, and be present in the greatest concentrations at the depth of one meter. Bifenthrin and chlorpyrifos are bound more strongly to the sediments, and were seen in smaller quantities at the specified depth.
In the last few decades, southeastern North Carolina has experienced an unprecedented population explosion, as the area around Wilmington becomes home to more people every year. As a consequence of this increase, the need for adequate recreational facilities has also been on the rise, as is evident by the large number of
new golf courses built in the area. As a result, there is always need for new and improved chemicals for the maintenance and general upkeep of the course. Of major environmental concern is the potential for course run-off entering the tidal creeks and estuarine systems during rain events, and upsetting the balance of the fragile ecosystem and the organisms that live there.
Some of southeastern North Carolina’s most important endangered natural resources are its wetlands and tidal creeks. These areas are important to many species of aquatic organisms which depend on these delicate areas for spawning and nurseries. In preliminary lab experiments, organic pesticides have been shown to exhibit toxic
effects on activity of acetylcholinesterase in the common fiddler crab, Uca pugilator1.
SPMDs: Semi-permeable membrane devices (SPMDs) were developed by Huckins et al.2 as passive in situ samplers containing purified triolein, a substance that constitutes the major portion of the lipid content of fish and other aquatic species. SPMDs are
normally constructed from low-density polyethylene tubing filled with high purity (95%)
triolein. The device is 91 cm long x 2.5 cm wide, with 1 gram of triolein lipid sealed inside the tubing. The SPMD is wrapped around a deployment cartridge called a spider, which serves as means for suspending the membrane inside the deployment
mechanism. Freely dissolved organics and other neutral substances diffuse through the
pores of the tubing, which have a diameter of approximately 10 angstroms, dissolving in the triolein, in much the same manner as a living fish gill membrane3. The lipid is extracted by dialysis and purified, and can then be analyzed by the traditional chemical procedures.
Accumulation of compounds with SPMDs is related to the individual compound’s
octanol-water partition coefficient (Kow), as well as its molecular size. The higher the value of the Kow, the more chemical will be sequestered by the triolein, and the longer it will take for equilibrium to be reached with the concentration in the water. Molecular size is a limiting factor of uptake, as large molecules will not be able to physically enter the
lipid layer through the pores. A major objective of this experiment was to access
the practicality of using SPMDs for the detection of trace levels of pesticide residues on golf courses. The Kow values for organic pesticides4 range from 100-107, which suggest high compatibility for use in conjunction with golf courses in determining the potential for movement off-site.
Experimental Methods:
Deployment of SPMD’s: Two SPMD’s were ordered from Environmental Sampling Technologies, St. Joseph, Missouri, and were shipped sealed in virgin paint cans along with a field blank. Both SPMD’s were deployed in the same device, which consisted of a
metal canister approximately eight-teen inches long with a diameter of about nine inches. The deployment device was perforated to allow water to pass through the canister and the SPMD’s. The deployment area was the mouth of Howe Creek, in a natural pool approximately four feet deep. Howe creek’s origins lie in Graham Pond, a man-made drainage lake built for the Jack Nicklaus golf course, and located about ten
feet from the deployment site. The trip blank was exposed to the surrounding air at 1:25 p.m. on May 5, 1998, and remained open during the time it took to load the deployment device with the two spiders containing the SPMD’s. Total time of exposure for the trip blank was 4 minutes, 57 seconds, measured from the time all cans were opened to the moment the deployment device was submerged in the creek. Deployment of SPMD’s,
and the start of the exposure period occurred at 1:30 p.m., May 5. Using stainless-steel clips and wire, the deployment device was lowered into the pool to a depth of about six-teen inches, and anchored to surrounding trees with the steel wire. The water temperature was measured at the time of deployment, and additional temperature reading were taken on May 7, 16, 23, 27 and June 1.
Spiders were recovered on June 1, 1998, during which the trip blank was re-exposed at 1:59 p.m. for a time of 1 minute, fifty seconds. There was a significant degree of biofouling of the membranes. The spiders were sealed in their original canisters, and transported immediately to the lab where they remained refrigerated until
their shipment to EST labs for clean up and dialysis.
Sample Processing: The SPMDs were shipped to EST labs in St. Joseph’s, Missouri for processing. The surface of each SPMD was cleaned to remove any debris accumulated during the exposure period. Each SPMD was dialyzed separately in
hexane at EST laboratories and sealed in its own vial, with the processed lipid dissolved
in 4 mls hexane. During the exposure period, one of the membranes was compromised and the triolein lipid contaminated with water. This lipid was concentrated and filtered through anhydrous sodium sulfate at EST in attempt to remove the water. After arrival at
UNCW, each sample was passed through a micro-column of silica with a small amount
of Na2SO4. This was done to remove any polar compounds that remain in the dialysate. The filtered sample was then concentrated to a volume of approximately 100 L with nitrogen gas. The sample was placed into a target vial for GC analysis.
Esterification of Samples: The initial GC spectra showed a substantial number of large peaks which were identified by GC/MS to be large fatty acids and hydrocarbon chains. These compounds can impair the sensitivity of the detector on the GC/MS and may obscure target compounds, and must be removed before any analysis can be attempted. These compounds were converted to more stable ester forms using a BF3-
methanol reagent. The concentrated sample was placed into a screw capped, micro
reaction vessel and allowed to equilibrate for twenty minutes at approximately 900C. A volume of 0.5 ml BF3-methanol reagent was added to the reaction vessel and allowed to reflux for an additional twenty minutes. The reaction temperature was not allowed to exceed 1000C. After twenty minutes, 1 ml of hexane/de-ionized H2O was added to the
vessel to stop the reaction. The vessel was agitated to extract the organics, and the
hexane layer removed. This layer was passed through an additional silica column, capped, and analyzed by GC and GC/MS. GC and GC/MS Analysis: The samples were initially analyzed using an SRI 8610 series
B gas chromatograph with a flame ionization detector. The GC was equipped with a 15 meter J.W. Scientific DB-5, 053mM Megabore column. The temperature program used for all GC analyses5 was as follows: initial temperature 800C, held for 1 minute, ramp 300C per minute to 1780C. Ramp 20C per minute to 2050C (13.5 minutes), then a final ramp of 300C per minute to 3000C. This high final temperature is to ensure complete
removal of all remaining samples from the column. 1 L of sample was injected without
split for each run. The GC/MS portion of this experiment was performed on a Hewlett-Packard 5890 series II gas chromatograph coupled with a HP 5971 mass selective detector. The
column used was a 15 meter J.W. Scientific DB-5. The mass scan parameters were set to the range of 45-350 atomic mass units. The temperature program remained the same throughout all sample runs. Any peaks of significant abundance were identified using the library data base NSB75k.
The SPMD extracts and field blank were analyzed using the FID to fine tune the temperature program and settings used for the GC/MS portion of the experiment. Three GC/MS runs were performed on the two samples, and one for the field blank. Chemical Movement Through Layered Soils (CMLS): CMLS was developed by
Nofzieger et al6, at Oklahoma State University for modeling chemical behavior in
homogeneous soils. The simulations can be used to predict the amount of chemical reaching a specified depth, and the time for this process to occur. Information about the weather, chemicals, site and soils to be used for the simulation can be entered into the database to be used by the computer for calculations about specific events. The
program must be supplied with soil organic content, bulk density, volumetric water
content (including wilting point, saturation, and field capacity) for each soil horizon. The organic partition coefficients and half-lives in soil must also be given for the compounds of interest. Using values for soil composition obtained from preliminary CMLS work at
UNCW7, and by obtaining the weather record for the month of May 1998, the simulator was run using five chemicals applied during the month of May, 1998. Information about the organic compounds applied to the Jack Nicholas golf course8 were obtained from the maintenance office at Landfall. This information allowed the utilization of the Extoxnet web site9 for acquiring the parameters for Half-life and soil partition
coefficients.
Using the edit function of the CMLS94 simulator, each category was completed with the new information for modeling. The three soil types used for this simulation were Murville fine sand (Mu), Seagate fine sand (Se), and Baymeade fine sand (Be). Each
soil has been shown to be present in area golf courses 7. The values entered for each sediment type were obtained from work conducted previously in the environmental lab at UNCW. Each replicate was obtained from the same site at 0.2 meter depth intervals. Each sample was entered as a separate horizon or layer in the soil edit function of the CMLS program.
Table 1: Soil characteristics used in the CMLS modeling program. The values are
reproduced from information previously obtained, where (Se) is Seagate fine sand, (Mu)
is Murville fine sand. And (Be) is Bayshore fine sand.
Sample # Bulk Density Field Capacity Wilting Point Saturation Organic Content
(g/cm^3) (%) (%) Point (%) (%)
Se1 1.096 6.70 2.70 58.63 3.50
Se2 1.204 4.07 1.07 54.55 0.57
Se3 1.425 4.34 1.24 46.22 0.72
Se4 1.327 4.87 1.47 49.93 0.60
Se5 1.421 8.61 2.41 46.37 2.20
Se6 1.308 7.12 1.02 50.66 2.59
Se7 1.402 4.60 1.10 47.09 1.34
Se8 1.448 4.42 0.92 45.37 1.10
Se9 1.450 4.63 1.13 45.28 0.69
Se10 1.612 4.40 0.90 39.17 0.59
Mu1 0.722 10.63 1.63 72.74 21.81
Mu2 0.994 13.50 4.50 62.47 10.26
Mu3 1.187 17.29 8.29 55.19 8.11
Mu4 1.353 15.69 6.69 48.93 4.63
Mu5 1.330 18.39 9.39 49.81 5.80
Mu6 1.133 16.62 7.62 57.24 6.16
Mu7 1.175 18.29 9.29 55.68 7.21
Mu8 1.289 17.20 10.20 51.35 5.90
Mu9 1.600 15.86 8.86 39.63 3.86
Be1 1.353 6.64 2.64 48.93 4.22
Be2 1.160 3.84 0.84 56.21 1.44
Be3 1.584 6.35 2.35 40.23 1.37
Be4 1.286 8.36 2.36 51.46 3.48
Be5 1.349 5.63 1.13 49.09 2.32
Be6 1.470 5.38 1.08 44.52 1.66
Be7 1.320 5.58 1.58 50.18 2.44
Be8 1.411 3.96 0.96 46.75 1.16
Be9 1.366 2.35 0.35 48.45 1.14
Be10 1.683 4.61 1.61 36.49 0.67
The chemicals were entered into the computer database using the edit function,
using values provided by Extoxnet for the organic partition coefficient and half-life. Table 2 also shows the amount of chemical applied to the course during the exposure period and each chemical’s water solubility in ppm. In order for the behavior of the five chemicals to be assessed accurately, an arbitrary amount of 50 kg/acre was entered as the initial amount applied in each simulation. This assures the amount of chemical
reaching the specified depth to be determined quantitatively. Table 2: Organic Compounds 8 applied to the Landfall course May 1998 for use with
CMLS simulations.
Chemical Trade Name Coeff. T1/2 9 Total Used (May 98) Water 9 Solubility
Bifenthrin Talstar 7.0 10 days 100 lbs. 0.1 ppm
Diquat Dibromide Reward -4.6 160 days 32 oz. 700,000 ppm @ 200C
Chlorpyrifos Dursban Pro 4.7 10 days 5 qt. 2 ppm @ 250C
Glyphosate Round-Up -3.1 47 days 4.5 oz. 12,000 ppm @
25 C
Fenoxycarb Award -4.3 47 days 7 lbs. 6 ppm @ 200C
The crop used for the simulation was wheat, since the database did not include values for the type of vegetation found on the course. Wheat has root system characteristics similar to grass, and could be expected to give similar results. The
weather data was obtained on-line, from the National Weather Service Wilmington
station (NWSILM) located at New Hanover National Airport. This information was recorded as an excel file, and then plugged into the program through the options screen. All experimental runs consisted of 250 simulations to insure the validity of the results. Runs were performed for each chemical with the three soil types. Statistics were
presented as a table under the heading Distribution of Travel Times and Amounts at a
fixed depth of one meter. The results were displayed as tables in the reports function, and converted to excel files from which figures depicting chemical movement and amount present were generated (see attached graphs).
Results and Discussion: In all cases, chromatogram spectra contained a high number of
large hydrocarbons, including various types of fatty acids which needed to be removed before analysis on the GC/MS could be conducted. Since these compounds disrupt the sensitivity of the mass detector on the GC/MS, they were esterified to remove the polar groups. There was also concern of these natural products obscuring peaks
corresponding to compounds of interest however no trace of pesticides, herbicides or
fungicides was detected after esterification. Figure 1 shows the original GC/MS run from the first SPMD vial, revealing the presence of numerous fatty acid peaks, such as oleic acid. The same sample is shown in Figure 2 after esterification, and the conversion to
ester derivatives is evident. It is also important to note the reduction in abundance, representing the removal of unwanted natural products sequestered by the SPMDs. Figure 3 shows similar characteristics, and the presence of butyl and methyl esters is
widespread in the spectrum. Analysis of the field blank by GC/MS showed only a
hexane solvent peak. Upon examination of these three spectra, no trace of any target species was found, although it is possible they are still obscured by the remaining natural products.
The CMLS program showed correlation with expected results based on half-lives
and partition coefficients of the chemicals used in the study. Diquat dibromide, fenoxycarb, and glyphosate demonstrated similar behavior in all soils, with near exact results obtained for travel times and amounts. This can be seen in the accompanying graphs for each soil. The most observations at one meter in Seagate and Baymeade
Fine Sand for these three compounds were at days 25 and 50, and at days 75 and 150
in Murville Fine Sand. This suggests that these compounds tend not to become bound to the sediments and may posses the ability to migrate off-site. However, due to the high solubility of diquat dibromide and glyphosate in water (7 x 105 and 1.2 x 104 respectively), it seems unlikely for the compounds to make it to the one meter target
depth, or be found in high concentrations in the any of the sediments. The amount of glyphosate, fenoxycarb, and diquat dibromide reaching one meter in Seagate was determined to be between 1- 10 kg/acre in all simulations. This correlates to a high percentage of the original compound, and is not an unreasonable prospect given the relatively long half-lives for the compounds (47, 47, and 160 days respectively). In
Murville, the amounts obtained were between 10-1 – 100 kg/acre for glyphosate and fenoxycarb, and 100- 101 kg/acre for diquat dibromide. The amounts found in Baymeade were 100- 101 kg/acre for glyphosate and fenoxycarb, and 101 – 10 2kg/acre for diquat dibromide. Taking this information into consideration, diquat dibromide seems to have the potential to do the most damage due to its penetration and persistence. The
travel times for bifenthrin and chlorpyrifos were similar at days 150 and 75 in Seagate and Baymeade soils. Data for bifenthrin in Murville Fine Sand were not available from this simulation. Chlorpyrifos reached one meter at days 150 and 200 days in Murville Fine Sand. This also is in agreement with the expected behavior due to the relatively large partition coefficient (4.7). These two chemicals would be expected to become
bound in the sediments, and not migrate off-site. The amounts for bifenthrin were found to lie between 10-2 – 10-1 kg/acre in Seagate, and 10 –3 – 10-2 in Baymeade. Chlorpyrifos yielded 10-1 - 100 in Seagate, 10-6 – 10-4 in Murville, and 10-4 – 10-3 kg/acre in Baymeade. This data is consistent with the short half-lives of these two compounds, and the simulations suggest that chlorpyrifos and bifenthrin would degrade faster,
leading to a lower concentration of the products in the sediments. Conclusions: The absence of organic species in the SPMD extracts was surprising due to the high Kow’s for these types of compounds. It is possible that the high volume in Graham Pond resulted in the concentrations of pollutants being below the detectable
limits of the instrumentation. Another possibility is the length of the exposure period may
have been insufficient to allow the accumulation of enough compounds for detection. In addition, the noted biofouling of the membrane after removal from the tidal creek could have resulted in decreased sequestering of contaminants, but it is still unclear how great a role this plays in the overall mechanism of SPMD uptake. The data obtained
from the SPMDs suggest that there is little danger of organic pesticides, herbicides, or
fungicides moving out of Graham retention pond, and posing a danger to the aquatic environment . However, the exposure period may have been too short to apply this statement to subsequent months, as these chemicals are used in various amounts year-
round.
The three chemicals found to move most readily through the soil types have variable effects on aquatic organisms. From the information provided by the Extoxonet web-site9, fenoxycarb is known to exhibit moderate to high toxicity towards fish, with an
LC50 ranging from 1.6 ppm in rainbow trout to 10.3 ppm in carp. Fenoxycarb is also
considered highly toxic to aquatic invertebrates such as Daphnia, with reproductive and growth effects evident after chronic exposures to concentrations greater than 1.6 ppb. However, it is interesting to note that when the pesticide was applied at rates ranging from 0.015 to 0.03 lbs/acre to ponds, no evidence of adverse effects were observed in
regards to invertebrates. In addition, the tissue residues of fish which had accumulated
20 times the ambient water concentration of pesticide quickly declined when the fish were removed from the test tanks, and placed in pesticide free waters. Therefore, it is unlikely that the compound would pose a threat to endangered species or organisms which consumed the fish. Diquat dibromide has an 8-hour LC50 of 12.3 ppm in rainbow
trout and 28.5 ppm in carp. However, studies indicate that yellow perch exhibit
significant respiratory stress when concentrations are similar to those used in aquatic vegetation control programs. Also, there appears to be no observed bioaccumulation of diquat dibromide in fish. Glyphosate is practically non toxic to fish, and may possibly be slightly toxic to invertebrates. The 96-hour LC50 is 120 ppm in bluegill sunfish, 168 ppm
in harlequin, and 86 ppm in rainbow trout. Reported values for other aquatic species
include 934 ppm for fiddler crabs, 281 ppm in shrimp, and 10 ppm in Atlantic oysters. Since there are several forms of the compound, certain formulations may exhibit a greater toxic effect than others, although it appears that in all cases, there is a very low potential for the compound to accumulate in body tissues or pose a threat to aquatic
species. The computer simulations show a limitation of the CMLS model with respect to sediment/water coupled systems. The program was intended for use in agricultural situations, and does not take the solubility’s of the chemicals in water into consideration
during computation of data. There are also several assumptions that must be made in order to perform the simulations. The model assumes the soil types used for the program exist at the target site, and in the same proportions. The wheat crop used in this model is not found on any golf courses in this country, and must be assumed to behave in relatively the same manner as the grasses found at the Landfall site. The
chemical database included in the program did not contain files for some relatively new chemicals currently in use on the Landfall course, and could only provide analysis of the environmental fate of certain chemicals, which may not necessarily be the most likely to be found in off-site locations. Therefore, it is difficult to take absolute confidence in the information provided by the simulator. Clearly, a more realistic model must be
implemented in order to assess the impact these organic species may have on the tidal creek environment. One future application of SPMDs involves the accumulation of pesticides by SPMDs in controlled laboratory experiments. Organophosphates have never been used
in conjunction with SPMDs, and much can be learned from the determination of their
sampling rates with respect to monitoring of water systems. The effect of temperature and salinity on SPMD sequestering ability is still not well understood, and is another candidate for future experimentation. The use of SPMDs for use in conjunction with golf
courses as monitoring devices has only begun, and further research in this area is
needed before their effectiveness can be accurately assessed.
Pesticide Analysis References 1. Cusac, Matthew; Effects of Acephate on Acetylcolinesterase Activities of Ribbed
Mussel Geukensia denissa and the Fiddler Crab Uca pugilator., 1998, Unpublished. 2. Huckins et al. Semipermeable membrane devices containing model lipid: A new approach to monitoring the bioavailability of lipophilic contaminants and estimating their bioconcentration potential. Chemoshpere, No. 20, 1990, pg. 533-553. 3. J.N Petty et al; Semipermeable membranes (SPMDs) for the concentration and
assessment of of bioavailable organic contaminants in aquatic environments.
Techniques in Aquatic Toxicology, Ostrander, G., Ed., Lewis Publishers, Boca Raton, 1995. 4. Rand, G., Ed. Fundamentals of Aquatic Toxicology, second edition; Taylor and Francis Publishing, New York, 1995; Table 1, pg. 498.
5. HP Environmental Solutions Catalog. 1995-96, page 78, EPA Method 507,
8140/8141. 6. Nofziger et al; Chemical Movement Through Layered Soils Modeling Program
(CMLS94), Version 95.09.18; Oklahoma State University, CMLS@SOILWATER.AGR.OKSTATE.EDU.
7. Ekberg, Eric; CMLS Soil and Research Paper; 1997, Unpublished.
8. Landfall Club Golf Maintenance Records, Pesticide use for January-July 1998; Wilmington, NC.
9. Extoxnet Web site; Extension Toxicology Network; http://www.maponus.com/ 10. Maps On Us; http://www.maponus.com/
15.0 Phosphorus, Cyanobacteria and Nitrogen Fixation in Stormwater
Detention Ponds
Lawrence B. Cahoon and Zhehong Ying Department of Biological Sciences UNC Wilmington
Introduction The primary function of stormwater detention ponds is to trap and retain certain pollutants, especially sediments. As sediments accumulate, ponds should tend to accumulate nutrients that are adsorbed by the sediments, particularly phosphorus.
Observations of local stormwater detention ponds show that they do actually trap
sediments and also tend to support higher biomasses of plants as they age, both of which will tend to trap nutrients, especially phosphorus compounds. Accumulation of phosphorus should tend to drive lower nitrogen:phosphorus (N:P) ratios, which should then favor growth of nitrogen-fixing cyanobacteria (blue-green algae). Observations over the period
1992-1998 indicate that cyanobacteria are very common in local stormwater detention
ponds, particularly ponds that have been in place for several years. Many of the cyanobacteria are filamentous forms that are usually strong nitrogen fixers. High rates of nitrogen fixation, if they are actually occurring, would then represent a significant source of "new" nitrogen to the drainages served by these detention ponds. Other observations
indicate strongly that nitrogen loading to the coastal waters downstream of local drainages
served by large numbers of detention ponds have been receiving elevated nitrogen inputs since about 1990. Therefore, we tested the following hypotheses in a variety of stormwater detention ponds: 1) phosphate accumulates in pond sediments as the ponds age, creating a phosphorus surplus and relative nitrogen deficiency, 2) phosphate accumulation favors
the growth of cyanobacteria, because they can fix their own nitrogen, thereby
circumventing nitrogen deficiency, and 3) older ponds therefore become a new source of fixed nitrogen. Methods
We surveyed a total of 14 ponds in New Hanover and Brunswick Counties with a range of ages and characteristics. All of these ponds serve a stormwater detention function, although some of them are natural ponds that have been altered for this purpose. Ages of the ponds were determined by reference to N.C. DENR records of stormwater pond permits or by local knowledge and records.
Sediment samples for analyses of total phosphate, microalgal biomass, and cyanobacterial biomass were collected from these ponds at 4 or 5 representative locations. Water samples were also collected and analyzed for total phosphate, transparency, and chlorophyll a (to calculate values of Carlson's Trophic State Index
[Carlson, 1977]). Total phosphate was determined by perchloric acid-nitric acid digestion for sediment samples, following Strickland and Parsons (1972), and by persulfate digestion for water column samples according to Valderrama (1981). Microalgal biomass was determined as chlorophyll a according to Whitney and Darley (1979), and water column chlorophyll a was determined fluorometrically according to Welschmeyer (1994).
Cyanobacterial biomass as phycoerythrin was determined using a fluorometric assay we developed in the laboratory (Ying, 1998).
Nitrogen fixation by water and sediment samples was estimated as acetylene
reduction to ethylene according to Flett et al. (1976), Kellar et al. (1977), and Cahoon and Kucklick (1993). Ethylene concentrations were measured on a gas chromatograph-mass spectrometer in the UNCW Chemistry Department against a Fisher ethylene gas standard.
Results The values of Carlson's Trophic State Index were all in a range corresponding to mesotrophic-eutrophic status and showed no relationship with pond age, indicating that water quality in these ponds very rapidly reached a stable condition, based on water clarity, total phosphate load, and chlorophyll a concentrations (Fig. 1).
Sediment phosphate contents of the ponds showed an increase with increasing pond age, although ponds approximately 30 years old were not much different from ponds approximately ten years old (Fig. 2). Sediment phosphorus contents varied considerably within each pond, probably because organic-rich and phosphorus-rich sediments tended
to accumulate more readily in some portions of each pond (often the deeper centers) than in others, such as the shallow fringes. The biomass of cyanobacteria in the water column and the sediments showed no increase with pond age, even exhibiting some decline with increasing age in the case of
water column cyanobacteria (Figs. 3 and 4). Cyanobacterial biomass was also not related in any clear way to total phosphorus in the water (Fig. 5) or in the sediments (Fig. 6). Nitrogen fixation rates, as measured by ethylene production, was negatively related or unrelated to total phosphate in the water column (Fig. 7) and in the sediments (Fig. 8).
Nitrogen fixation rates assessed under light and dark conditions appeared to demonstrate fixation by heterotrophic bacteria in addition to phototrophic cyanobacteria (Figs. 7 and 8). Discussion The data presented here support the hypothesis that stormwater detention ponds
accumulate phosphorus with age, a trapping function that can improve water quality
downstream by reducing phosphorus loading. Sediment phosphorus concentrations
appeared to plateau at approximately 200 µg/g sediment, a level about twice as high as phosphorus-enriched sediments from Lake Waccamaw (Cahoon et al., 1990; Cahoon and
Kucklick, 1993). The leveling of sediment phosphorus concentrations after a few years
suggests, however, that pond sediments are rapidly saturated with phosphorus, then fail to retain more as it enters, suggesting either removal by increasing plant growth or export during flushing events. Since plant growth is usually very seasonal, periods of slow plant growth or dieback may create periods of higher than average rates of export of
phosphorus from detention ponds.
The data do not support the hypothesis that cyanobacterial biomass increases as ponds age and trap more phosphorus. There are several explanations consistent with this conclusion. First, it is possible that cyanobacteria become more closely associated with
macrophytes as ponds age and macrophyte stands develop, so that sampling of the open
water and sediments missed the cyanobacteria and their nitrogen fixation activity.
Quantitative macrophyte sampling is difficult and there is no "standard" method for it, so sampling of cyanobacterial periphyton is similarly difficult. Second, there is a possibility that nitrogen fixing heterotrophic bacteria become more important than cyanobacterial
nitrogen fixers in some ponds at some times. We made no effort to distinguish these two
somewhat differently behaved groups of nitrogen fixers. Third, the premise of the phosphate overloading hypothesis is that a surplus of resident phosphorus would create a nitrogen deficiency that would favor cyanobacteria that fix nitrogen. However, the rapid turnover time of stormwater detention ponds, which are designed to hold water during high
flows for very short periods, might supply sufficient nitrogen via inflows to limit times of
nitrogen deficiency to low flow periods. The data also do not support the hypothesis that nitrogen fixation rates are higher in older, more phosphorus-enriched ponds. The same reasons advanced above to explain
the patterns of cyanobacterial biomass distribution may be invoked as explanations for this
conclusion. Nevertheless, nitrogen fixation rates observed in many of the ponds are not trivial, and represent rates of nitrogen loading that may be quite significant. For example, the higher nitrogen fixation rates shown for some ponds in Figs. 7 and 8 are very similar to nitrogen fixation rates reported from Lake Waccamaw, where nitrogen fixation is
considered to be a significant source of "new" nitrogen (Cahoon and Kucklick, 1993).
However, the relatively high throughput rates of water in stormwater detention ponds likely make nitrogen inputs from runoff a much greater part of the overall nitrogen budget than inputs from nitrogen fixation. At this time, this is an issue that would require additional study to resolve.
In summary, stormwater detention ponds probably trap nutrients in a more complex ways than they trap sediment. There is evidence from this study that detention ponds retain and trap phosphorus but reach a saturation level, after which they probably release as much phosphorus as they take in. However, the role of aquatic plants in temporarily
sequestering phosphorus during the growing season must be considered. High rates of
nutrient loading through runoff are likely balanced by high exports of soluble nitrogen, but may drive lower nitrogen fixation rates than would be expected simply from static measures of N:P ratios or phosphorus concentrations. Nitrogen fixation may rise when water residence times are long, and may also be rapid when high biomasses of
cyanobacteria are associated with aquatic macrophytes. Thus, the role of aquatic
macrophytes must be viewed as a central part of nutrient cycling and management in stormwater detention ponds.
16.0 References Cited APHA. 1995. Standard Methods for the Examination of Water and Wastewater, 19th ed.
American Public Health Association, Washington, D.C.
Burkholder, J.M., H.B. Glasgow, Jr. and C.W. Hobbs. 1995. Fish kills linked to a toxic ambush-predator dinoflagellate: distribution and environmental conditions. Marine Ecology Progress Series 124:43-61.
Burkholder, J.M. and H.B. Glasgow, Jr. 1997. Pfiesteria piscicida and other Pfiesteria-like dinoflagellates: Behavior, impacts, and environmental controls. Limnology and Oceanography 42:1052-1075.
Cahoon, L.B., and J.R. Kucklick. 1993. Characteristics of nitrogen fixation in Lake
Waccamaw, North Carolina. J. Elisha Mitchell Sci. Soc. 109:20-29. Cahoon, L.B., J.R. Kucklick, and J.C. Stager. 1990. A natural phosphate source for Lake Waccamaw, North Carolina, USA. Int. Rev. ges. Hydrobiol. 75:339-351.
Carlson, R.E. 1977. A trophic state index of lakes. Limnology and Oceanography 22:361-369. Flett, R.J., R.D. Hamilton, and N.E.R. Campbell. 1976. Aquatic acetylene-reduction
techniques: solutions to several problems. Canadian Journal of Microbiology 22:43-51.
Hales, J. 1998. Tidal flushing in coastal estuaries: Futch Creek. Unpublished manuscript, Chemistry Department, UNC Wilmington. Kellar, P.E., S.A. Paulson, and L.J. Paulson. 1977. Methods for biological, chemical and
physical analyses in reservoirs. Lake Mead Limnological Research Center Technical
Report Series 5:55-76. Long, E.R., D.D. McDonald, 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.
Mallin, M.A., L.B. Cahoon, E.C. Esham, J.J. Manock, J.F. Merritt, M.H. Posey, T.D. Alphin, R.K. Sizemore and K. Williams. 1996. Water Quality in New Hanover County
Tidal Creeks, 1995-1996. Center for Marine Science Research, University of North Carolina at Wilmington, Wilmington, N.C.
Mallin, M.A., M.H. Posey, M.L. Moser, G.C. Shank, M.R. McIver, T.D. Alphin, S.H. Ensign and J.F. Merritt. 1998. Environmental Assessment of the Lower Cape Fear
River System, 1997-1998. CMSR Report No. 98-02, Center for Marine Science Research, University of North Carolina at Wilmington, Wilmington, N.C.
Mallin, M.A. and T.L. Wheeler. 1998. Nutrient and fecal coliform discharge from
southeastern North Carolina golf courses. pp 22-25 In Is Golfing Greener? - The Impact of Golf Courses on the Environment. Symposium Proceedings, The University
of North Carolina at Wilmington/ NC Cooperative Extension Service/ NC Coastal
Federation, March 12, 1998. Parsons, T.R., Y. Maita and C.M. Lalli. 1984. A Manual of Chemical and Biological Methods for Seawater Analysis. Pergamon Press, Oxford. 173 pp.
Rhoads, J.M., 1998. Utilization of isolated and mixed structural habitats in an estuarine system. M.S. Thesis, University of North Carolina at Wilmington, Wilmington, N.C. Sargent, W.B. and Carlson, P.R., Jr., 1987. The utilization of Breder traps for sampling
mangrove and high marsh fish assemblages. In: Proceedings of the Fourteenth Annual Conference on Wetlands Restoration and Creation. Webb, F. (ed.) Hillsborough Community College, Tampa, Fl. p.194-205. Schlotzhauer, S.D. and R.C. Littell. 1987. SAS system for elementary statistical
analysis. SAS Institute, Inc., SAS Campus Dr., Cary, N.C. Strickland, J.D.H., and T.R. Parsons. 1972. A Practical Handbook of Seawater Analysis. Fisheries Research Board of Canada, Ottawa.
Townsend, E.C. 1991. Depth distribution of the grass shrimp Palaemonetes pugio into contrasting tidal creeks in North Carolina and Maryland. M.S. Thesis, University of North Carolina at Wilmington, Wilmington, N.C. U.S. EPA 1997. Methods for the Determination of Chemical Substances in Marine and
Estuarine Environmental Matrices, 2nd Ed. EPA/600/R-97/072. National Exposure Research Laboratory, Office of Research and Development, U.S. Environmental Protection Agency, Cincinnati, Ohio. Valderrama, J.C. 1981. The simultaneous analysis of total nitrogen and phosphorus in
natural waters. Marine Chemistry 10:109-122. Walker, W.J. and B. Branham. 1992. Environmental impacts of turfgrass fertilization. p 105-219. In J.C. Balogh and W.J. Walker, (eds.) Golf Course Management and Construction: Environmental Issues. Lewis Publishers, Chelsea, MI.
Webb, F.J., 1987. Proceedings of the Fourteenth Annual Conference of Wetlands Restoration and Creation. Hillsborough Community College, Tampa, Fl. 157p. Welschmeyer, N.A. 1994. Fluorometric analysis of chlorophyll a in the presence of
chlorophyll b and phaeopigments. Limnology and Oceanography 39:1985-1993. Whitney, D.E., and W.M. Darley. 1979. A method for the determination of chlorophyll a in samples containing degradation products. Limnology and Oceanography 24:183-187.
Ying, Z. 1998. Phosphorus, cyanobacteria and nitrogen fixation in stormwater detention ponds. Unpublished M.S. thesis, UNC Wilmington, Wilmington, N.C.
17.0 Acknowledgments Funding for this research was provided by the City of Wilmington Engineering
Department; the New Hanover County; the Northeast New Hanover Conservancy;
and North Carolina State University. For project facilitation and helpful information we thank Paul Foster, Dr. Christopher Halkedies, Dexter Hayes, Patrick Lowe, David Mayes, Karen Neutzling, Dr. Pam Seaton and Dave Weaver. Pfiesteria counts were performed by Dr. JoAnn Burkholder of North Carolina State University. For field and
laboratory assistance we thank Scott Ensign, Mark Gay, Christopher Jolly, Matt
McIver, Christian Preziosi, Chris Shank, Ashley Skeen, and Mark Wainright.
18.0 Appendix A. Guideline values for sediment metals concentrations (ppm, dry wt.) potentially harmful to aquatic life (Long et al. 1995). _____________________________________________________________________
ERL (Effects range low) concentrations below ERL are those in which harmful effects
are rarely observed. ERM (Effects range median) concentrations above ERM are those in which harmful effects would frequently occur. Concentrations between ERL and ERM are those in which harmful effects would
occasionally occur.
Chemical ERL ERM
Arsenic 8.2 70.0
Cadmium 1.2 9.6
Chromium 81.0 370.0 Copper 34.0 270.0 Lead 46.7 218.0 Mercury 0.15 0.71
Nickel 20.9 51.6 Silver 1.0 3.7 Zinc 150.0 410.0
Figure 1.1 Futch, Pages, Howe, Bradley, Hewletts, and Whiskey Creeks in New Hanover County, North Carolina, USA.