NOAA Technical Memorandum NOS NCCOS CCMA 137

National Status and Trends Program

for Marine Environmental Quality

Survey of Sediment Quality in

Sabine Lake, Texas and Vicinity

Silver Spring, Maryland

December 1999

US Department of Commerce

noaa National Oceanic and Atmospheric Administration

Center for Coastal Monitoring and Assessment

National Centers for Coastal Ocean Science

National Ocean Service

Sabine Lake

Sabine Pass

Gulf of Mexico

Neches River

Beaumont

Sabine River

Port 

Neches

Sabine-Neches Canal

Texas

Talor

 Basin

Port

 Arthur

Louisiana

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0

1000

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5000

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total PAHs, 

ppb

ERL = 

4022
Center for Coastal Monitoring and Assessment

National Centers for Coastal Ocean Science

National Ocean Service

National Oceanic and Atmospheric Administration

U.S. Department of Commerce

N/ORCA2, SSMC4

1305 East-West Highway

Silver Spring, MD 20910

Notice

This report has been reviewed by the National Ocean Service of the National Oceanic and

Atmospheric Administration (NOAA) and approved for publication. Such approval does not

signify that the contents of this report necessarily represents the official position of NOAA or

of the Government of the United States, nor does mention of trade names or commerical

products constitute endorsement or recommendation for their use.
NOAA Technical Memorandum NOS ORCA 137

Survey of Sediment Quality in

Sabine Lake, Texas and Vicinity

Edward R. Long

National Oceanic and Atmospheric Administration

Contributors:

M. Jawed Hameedi, Michelle Harmon

National Oceanic and Atmospheric Administration

Gail M. Sloane

Florida Department of Environmental Protection

R. Scott Carr, James Biedenbach, Tom Johnson

U. S. Geological Survey

K. John Scott, Cornelia Mueller

Science Applications International Corporation

Jack W. Anderson

Columbia Analytical Services, Inc.

Terry L. Wade, Bobby J. Presley

Texas A & M University

Silver Spring, Maryland

December 1999

United States National Oceanic and National Ocean Service

Department of Commerce Atmospheric Administration

William M. Daley D. James Baker Nancy Foster

Secretary Under Secretary Assistant Administrator
Contents

Abstract ........................................................................................................................................... 1

Introduction .................................................................................................................................... 2

Background ............................................................................................................................. 2

Goals and Objectives ............................................................................................................. 2

Methods........................................................................................................................................... 3

Sampling Design .................................................................................................................... 3

Laboratory Toxicity Tests ...................................................................................................... 7

Cytochrome p-450 RGS Assays ............................................................................................ 12

Chemical Analyses—Metals ................................................................................................. 12

Chemical Anayses—Organic Compounds ......................................................................... 13

Chemistry QA-QC.................................................................................................................. 14

Benthic Community ............................................................................................................... 15

Statistical Methods ................................................................................................................. 15

Results ............................................................................................................................................. 18

Severity and Spatial Patterns in Toxicity ............................................................................ 18

Spatial Extent of Toxicity ....................................................................................................... 22

Summary of Toxicity Results ................................................................................................ 28

Spatial Patterns in Chemical Concentrations ..................................................................... 28

Spatial Extent of Chemical Contamination ........................................................................ 31

Relationships Between Toxicity and Chemical Concentrations ...................................... 31

Summary of Chemical Contamination ............................................................................... 34

Discussion and Conclusions ........................................................................................................ 34

Acknowledgements ....................................................................................................................... 39

References ....................................................................................................................................... 39

Appendix A .................................................................................................................................... 42

Tables

1. Trace Metals Measured in Sabine Lake Sediments and Method Detection Limits .......... 13

2. Organic Compounds Measured in Sabine Lake Sediments and Method

Detection Limits ............................................................................................................................. 14

3. Summary of Toxicity Test Results for 66 Sediment Sampling Stations in

Sabine Lake ..................................................................................................................................... 19

4. Estimates of the Spatial Extent of Toxicity in Sediments of Sabine Lake Based Upon

Results from 5 Independent Tests ............................................................................................... 27
5. Estimates of the Spatial Extent of Chemical Contamination Relative to

Effects-Based Numerical Guidelines (ERL and ERM) for Sabine Lake ................................. 32

6. Spearman-Rank Correlation Coeffecients (rho, corrected for ties) for Toxicity

Test Results and Chemical Concentrations Normalized to ERM Values in

Sabine Lake Sediments ................................................................................................................. 32

7. Spearman-Rank Correlation Coeffecients (rho, corrected for ties) for

P-450 RGS Test Results and Concentrations of Classes of Organic

Compounds in Sabine Lake Sediments ...................................................................................... 33

8. Spearman-Rank Correlation Coeffecients (rho, corrected for ties) for

Percent UrchinNormal Development (100% Porewater) and Concentrations of

Trace Metals and Unionized Ammonia in Sabine Lake Sediments........................................ 33

9. A Comparison of the Spatial Extent of Toxicity (Percentage of Total Survey

Area) in Amphipod Survival Tests Among 25 Estaurine Areas Nationwide and

“National Estuarine Averages Compiled from Field Studies Conducted Through

1995, 1996, and 1997” .................................................................................................................... 35

10. A Comparison of the Spatial Extent of Toxicity (Percentage of Total Survey

Area) in Urchin Fertilization Tests of 100% Porewater Among 22 Estuarine

Survey Areasand “National Estuarine Averages Compiled from Field Studies

Conducted Through 1995, 1996, and 1997” ............................................................................... 36

11. A Comparison of the Spatial Extent of Toxicity (Percentages of Area) in Urchin

Embryo Development Tests Performed with 100% of Porewater of Sediments

from 8 U.S. Bays and Estuaries .................................................................................................... 37

12. Spatial Extent of Toxicity (KM2 and Percentages of Total Area) in Microbial

Bioluminescence Tests Performed with Solvent Extracts of Sediments from 17 U.S.

Bays and Estuaries ......................................................................................................................... 38

13. Spatial Extent of Toxicity (KM2 and Percentages of Total Area) in P-450 RGS

Bioassays Performed with Solvent Extracts of Sediments from 6 Areas ............................... 38

Figures

1. Sabine Lake and Surrounding Study Area............................................................................. 4

2. Locations and Dimensions of Sampling Strata and Substrata in Sabine Lake ................. 5

3. Locations of Individual Sampling Stations within each Stratum for Sabine Lake........... 6
4. Sampling Stations in which Results of Amphipod Survival Tests were not

Significant, Significant Only, or Highly Significant .......................................................... 21

5. Sampling Stations in which Results of Amphipod Survival Tests were either

not Significant, or Significant in 100%, 50%, or 25% Porewater Concentrations .......... 23

6. Sampling Stations in which Results of Amphipod Survival Tests were either not

Significant, or Significant (p<0.05) in 100%, 50%, or 25% Porewater Concentrations . 24

7. Results of Microtox Tests Expressed as Ratio of Sample Mean to Reference Mean......... 25

8. Results of Cytochrome P-450 RGS Assays Expressed as Benzo(a)Pyrene Equivalents... 26

9. Concentrations of Arsenic in Sediments from Sabine Lake and Vicinity .......................... 29

10. Concentrations of Total PAHs in Sediments from Sabine Lake and Vicinity ................. 30
1

ABSTRACT

The toxicity of sediments in Sabine Lake, Texas, and adjoining Intracoastal Waterway

canals was determined as part of bioeffects assessment studies managed by NOAA’s

National Status and Trends Program. The objectives of the survey were to determine: (1)

the incidence and degree of toxicity of sediments throughout the study area; (2) the spatial

patterns (or gradients) in chemical contamination and toxicity, if any, throughout the study

area; (3) the spatial extent of chemical contamination and toxicity; and (4) the statistical

relationships between measures of toxicity and concentrations of chemicals in the sediments.

Surficial sediment samples were collected during August, 1995 from 66 randomly-chosen

locations. Laboratory toxicity tests were performed as indicators of potential

ecotoxicological effects in sediments. A battery of tests was performed to generate information from different phases (components) of the sediments. Tests were selected to represent a range in toxicological endpoints from acute to chronic sublethal responses. Toxicological tests were conducted to measure: reduced survival of adult amphipods exposed to

solid-phase sediments; impaired fertilization success and abnormal morphological development in gametes and embryos, respectively, of sea urchins exposed to pore waters; reduced metabolic activity of a marine bioluminescent bacteria exposed to organic solvent

extracts; and induction of a cytochrome P-450 reporter gene system in exposures to solvent extracts of the sediments.

Chemical analyses were performed on portions of each sample to quantify the concentrations of trace metals, polynuclear aromatic hydrocarbons, and chlorinated organic compounds. Correlation analyses were conducted to determine the relationships between

measures of toxicity and concentrations of potentially toxic substances in the samples.

Based upon the compilation of results from chemical analyses and toxicity tests, the quality

of sediments in Sabine Lake and vicinity did not appear to be severely degraded. Chemical

concentrations rarely exceeded effects-based numerical guidelines, suggesting that toxicant-induced effects would not be expected in most areas. None of the samples was highly

toxic in acute amphipod survival tests and a minority (23%) of samples were highly toxic in

sublethal urchin fertilization tests. Although toxic responses occurred frequently (94% of

samples) in urchin embryo development tests performed with 100% pore waters, toxicity

diminished markedly in tests done with diluted pore waters. Microbial bioluminescent

activity was not reduced to a great degree (no EC50 <0.06 mg/ml) and cytochrome P-450

activity was not highly induced (6 samples exceeded 37.1 ug/g benzo[a]pyrene equivalents) in tests done with organic solvent extracts. Urchin embryological development was

highly correlated with concentrations of ammonia and many trace metals. Cytochrome P450 induction was highly correlated with concentrations of a number of classes of organic

compounds (including the polynuclear aromatic hydrocarbons and chlorinated compounds).
2

INTRODUCTION

Background.

As a part of the National Status and Trends (NS&T) Program, NOAA conducts assessments of the adverse biological effects of toxic chemicals in selected coastal regions and

estuaries. Studies are performed to determine biological effects of toxicants in fishes and

bivalve molluscs. This report is one in a series of regional reports on sediment quality.

Previous reports have been produced for the Hudson-Raritan estuary, Long Island Sound,

Boston Harbor, Tampa Bay, San Diego Bay, San Pedro Bay, southern California estuaries,

western Florida panhandle, and South Carolina/Georgia bays and summarized in Long et

al. (1996).

Sabine Lake is an inland estuary that straddles the Texas/Louisiana border near Beaumont,

Texas. The estuary is adjacent to a portion of the Intracoastal Waterway (ICW), which is

composed of several channels of the Neches and Sabine rivers. Freshwater is supplied to

the lake by these two rivers. The lake is open to the Gulf of Mexico through Sabine Pass.

Trace metals concentrations in sediments have been reported as relatively low (even depleted relative to other nearby areas) in the Sabine-Neches estuary as compared to other

estuaries along the Gulf coast (Ravichandran et al., 1995). However, Harrell and

McConnell (1995) reported detectable concentrations of dioxins and furans in the clam

Rangia cuneata in the Neches River downstream from a pulp mill. There is a huge complex of petroleum – related industries along the Neches and Sabine rivers, particularly

many refineries and transhipment docks near Beaumont and Port Arthur. Crude oil and

petroleum products are shipped and piped on- and offshore in this area.

Sabine Lake was chosen for this survey based upon the likelihood of chemical contamination within sediments of the area and an interest expressed by the State of Texas. The

survey of Sabine Lake was followed in 1996 by a similar survey of Galveston Bay; the

results of which are reported in a separate report.

Goals and Objectives.

The overall goal of this study was to provide a characterization of the toxicological condition

of sediments in Sabine Lake and surrounding channels, as a measure, or indicator, of

adverse biological effects of toxic chemicals. Data from toxicity tests were intended to

provide a means of determining whether toxic conditions actually occurred within any of the

areas.

Several specific technical objectives were established to serve as guides for the sampling

designs and analytical plans. The objectives of the study were to:

(1) determine the incidence and degree of toxicity of sediments throughout the study area;

(2) determine the spatial patterns (or gradients) in chemical contamination and toxicity

throughout the study area;

(3) determine the spatial (or surficial) extent of chemical contamination and toxicity as

percentage of the total study area; and
3

(4) determine the statistical relationships between measures of toxicity, concentrations of

chemical substances and benthic populations in the sediments.

METHODS

Sampling design.

The study area encompassed all of Sabine Lake, portions of the Sabine River, portions of

the Neches River, portions of the Neches-Sabine Canal at the confluence of the two rivers,

portions of Sabine Pass channel entrance, and an area in the Gulf of Mexico near the

entrance channel (Figure 1). A stratified-random sampling design similar to those used in

previous surveys conducted nationwide by NOAA (Long et al., 1996) was applied in Sabine

Lake. The study area was subdivided into 22 irregular-shaped strata, designated A through

R (Figure 2). Strata established within channels were further subdivided into three substrata to improve spatial coverage. Only one location each was sampled within each substratum, whereas three locations were sampled in each of the larger undivided strata (i.e.,

strata M, R, and P).

This approach provided the least intense sampling effort in areas known or suspected to be

relatively homogeneous in sediment type and water depth, and relatively distant from

contaminant sources. In contrast, relatively small strata were established in channels and

urban harbors nearer suspected sources in which conditions were expected to be heterogeneous or transitional. As a result, sampling effort was more intense in the small strata

than in the large strata. Strata R and P were established to document conditions beyond

the mouth of the Sabine Pass channel.

This approach combines the strengths of a stratified design with the random-probabilistic

selection of sampling locations. Data generated within each stratum or sub-stratum can be

attributed to the dimensions of that area. Therefore, these data can be used to estimate

the spatial extent of toxicity with a quantifiable degree of confidence (Heimbuch, et al.,

1995). Strata boundaries were established to coincide with the dimensions of major basins, bays, inlets, waterways, etc. in which hydrographic, bathymetric and sedimentological

conditions were expected to be relatively homogeneous. Sub-stratum boundaries were

established at roughly equal intervals along the axes of the channels.

Within the boundaries of each stratum, all possible latitude/longitude intersections had an

equal probability of being selected as a sampling location. The locations of individual

sampling stations within each stratum were chosen randomly using GINPRO software,

developed by NOAA, applied to a digitized navigation chart. Four alternate locations were

provided for each station in a numbered sequence. The coordinates for each alternate

were provided in tables and were plotted on the appropriate navigation chart. In two cases

the coordinates provided were inaccessible (first alternate for station 3 blocked by a dredging operation) or only shell and cobble were present at the location (oyster shells in first

alternate for station 57). In those cases, the first station alternate was rejected and the

vessel was moved to the second alternate set of coordinates to collect the sample. In both

cases, the second location was accessible and sediments were acceptable.
4

Sabine Lake

Sabine Pass

Gulf of Mexico

Neches River

Beaumont

Sabine River

Port 

Neches

Sabine-Neches Canal

Texas

Talor

 Basin

Port

 Arthur

Louisiana

igure 1. Sabine Lake and surrounding study area.
5

L-1

L-2

L-3

M-1

M-2

M-3

M-5 M-4

M-6

M-7

O-1

O-2

O-3

 R

 P

G-1

G-2

G-3

H-1

H-2

H-3

I-1

I-2

I-3

J-1 J-2

J-3

K-1

K-2

K-3

B-1 B-2

D-3 A-3

D-2

D-1

A-1

A-2

B-3

C-1 C-2

C-3 F-3

F-2 F-1

E-3

E-2 E-1

Figure 2. Locations and dimensions of sampling strata and sub-strata in Sabine Lake.
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Figure 3. Locations of individual sampling stations within each stratum for Sabine

Lake.
7

Sediments from a total of 66 stations were collected during August, 1995 with the NOAA

vessel Ferrel and its launch. Each station was sampled only once.

Vessel positioning and navigation were aided with a differential-corrected, Global Positioning System (GPS) unit and a compensated LORAN C unit. The two systems generally

agreed well when both were operational.

Samples for toxicity and chemical testing were collected with a Kynar-lined 0.1m2 modified

van Veen grab sampler (also, known as a Young grab) deployed with a hydraulic winch.

The grab sampler and sampling utensils were acid washed with 10% HCl at the beginning

of each day, and thoroughly cleaned with site water and acetone before each sample

collection. Usually, 3 or 4 deployments of the sampler were required to provide a sufficient

volume of material for the toxicity tests and chemical analyses. The upper 2-3 cm of the

sediment were sampled with a plastic medical scoop and accumulated in a stainless steel

pot. The pot was covered with a Teflon plate between deployments of the sampler to minimize sample oxidation and exposure to shipboard contamination. The material was carefully homogenized in the field with a stainless steel spoon before it was distributed to prepared containers for shipment to respective testing laboratories.

Samples for benthic community analyses were collected at one station randomly chosen

within each stratum. Triplicate samples were collected at each station with a Young-modified, petite (0.413 cm2

) van Veen grab. Samples for both toxicity/chemistry analyses and

the benthic community analyses were collected at the same location. The entire contents

of samples that were at least 5 cm deep were retained and sieved in the field with a 0.5 mm

sieve. Material retained on the sieve was preserved in 10% buffered formalin with Rose

Bengal. Samples were rejected if the jaws of the grab were open, if the sample was partly

washed out or if the sample was less than 5 cm deep. A fourth sample was collected at

each location and material retained for total organic carbon and grain size analyses.

Sample jars for each toxicity test and chemistry analysis were sealed to prevent leakage

and outside contamination and shipped in ice chests packed with frozen water bottles or

blue ice to the testing laboratories by overnight courier. Samples for toxicity tests were kept

chilled until extractions or tests were initiated. Samples for chemical analyses and cytochrome P-450 tests were kept frozen until thawed for analyses. All samples were accompanied by chain of custody forms which included the date and time of the sample collection

and station designation.

Locations, water depths, sampling dates, and visual sediment characteristics of the individual sampling stations for each sampling stratum are summarized in Appendix A. Multiple

toxicity tests and complete chemical analyses were performed on all 66 sediment samples.

Laboratory toxicity tests.

Multiple toxicity tests were performed on aliquots of each sample to provide a weight of

evidence. Tests were selected for which there were widely-accepted protocols, that would

represent the toxicological conditions within different phases (partitions) of the sediments,

and that would likely represent a range in response from physiological to acute lethal endpoints.
8

Amphipod survival test. The amphipod tests are the most widely and frequently used

assays in sediment evaluations performed in North America. They are performed with

subadult crustaceans exposed to relatively unaltered, bulk sediments. Ampelisca abdita

has shown relatively little sensitivity to factors such as grain size, ammonia, and organic

carbon in previous surveys. In previous surveys performed by the NS&T Program (Long et

al., 1996), this test has provided wide ranges in responses among samples, strong statistical associations with elevated toxicant levels, and small within-sample variability.

Ampelisca abdita is a euryhaline benthic amphipod that ranges from Newfoundland to

south-central Florida, and along the eastern Gulf of Mexico. Also, it is abundant in San

Francisco Bay along the Pacific coast. The amphipod test with A. abdita has been routinely

used for sediment toxicity tests in support of numerous EPA programs, including the Environmental Monitoring and Assessment Program (EMAP) in the Virginian, Louisianian,

Californian, and Carolinian provinces (Schimmel et al., 1994).

Amphipod survival tests were conducted by Science Applications International Corporation,

(SAIC) in Narragansett, R.I. All tests were initiated within 10 days of the date samples were

collected. Samples were shipped by overnight courier in 4-liter high density polyethylene

jugs which had been washed, acid-stripped, and rinsed with deionized water. Sample jugs

were packed in shipping coolers with blue ice. Each was inspected to ensure it was within

acceptable temperature limits upon arrival and then the jugs were stored at 4°C until testing

was initiated. Prior to testing, sediments were mixed with a stainless steel paddle and

press-sieved through a 1.0-mm mesh sieve to remove debris, stones, resident biota, etc.

Amphipods were collected by SAIC from tidal flats in the Pettaquamscutt (Narrow) River, a

small estuary flowing into Narragansett Bay, RI. Animals were held in the laboratory in presieved uncontaminated (“home”) sediments under static conditions. Fifty percent of the

water in the holding containers was replaced every second day when the amphipods were

fed. During holding, A. abdita were fed laboratory cultured diatoms (Phaeodactylum

tricornutum). Control sediments were collected by SAIC from the Central Long Island

Sound (CLIS) reference station of the U.S Army Corps of Engineers, New England Division.

These sediments have been tested repeatedly with the amphipod survival test and other

assays and found to be nontoxic (amphipod survival has exceeded 90% in 85% of the

tests) and uncontaminated (Long et al., 1996). Sub-samples of the CLIS sediments were

tested along with each series of samples from northern Puget Sound.

Amphipod testing followed the procedures detailed in the Standard Guide for conducting

10-day Static Sediment Toxicity Tests with Marine and Estuarine Amphipods (ASTM, 1992).

Briefly, amphipods were exposed to test and negative control sediments for 10 days with 5

replicates of 20 animals each under static conditions using filtered seawater. Aliquots of

200 mls of test or control sediments were placed in the bottom of one-liter test chambers,

and covered with approximately 600 mL of filtered seawater (28-30 ppt). Air was provided

by air pumps and delivered into the water column through a pipette to ensure acceptable

oxygen concentrations. Pipettes were suspended in a manner to ensure that the sediments would not be disturbed.

Temperature was maintained at ~20°C by a temperature-controlled water bath. Lighting

was continuous during the 10 day exposure period to inhibit emergence of the organisms
9

from the sediment, thereby maximizing the exposure of the amphipods to the test sediments. Information on temperature, salinity, dissolved oxygen, pH and ammonia in the test

chambers was obtained during tests of each batch of samples to ensure compliance within

acceptable ranges. Ammonia concentrations were determined in both pore waters (day 0

of the tests) and overlying waters (days 2 and 8 of the tests). Concentrations of the unionized form of ammonia were calculated, based upon measures of total ammonia and concurrent measures of pH, salinity and temperature.

Twenty healthy, active animals were placed in each test chamber, and monitored to ensure

they burrowed into sediments. Non-burrowing animals were replaced. The jars were

checked daily, and records kept of dead animals, and animals on the water surface,

emerged on the sediment surface, or in the water column. Those on the water surface

were gently freed from the surface film to enable them to burrow, and dead amphipods

were removed.

Tests were terminated after ten days. Contents of each of the test chambers were sieved

through a 0.5 mm mesh screen. The animals and any other material retained on the

screen were examined under a stereomicroscope for the presence of amphipods. Total

amphipod mortality was recorded for each test replicate.

A positive control (reference toxicant) test was used to document the sensitivity of each

batch of test organisms. The positive control consisted of 96 hr water-only exposures to

sodium dodecyl sulfate (SDS). LC50 values were calculated for each test run using results

from tests of five SDS concentrations. Control charts provided by SAIC showed consistent

results in tests of both the positive and negative controls.

Sea urchin fertilization test. Tests of sea urchin fertilization have been used in assessments of ambient water and effluents and in previous NS&T Program surveys of sediment

toxicity (Long et al., 1996). Test results have shown wide ranges in responses among test

samples, excellent within-sample homogeneity, and strong associations with the concentrations of toxicants in the sediments. This test combines the features of testing sediment pore

waters (the phase of sediments in which dissolved toxicants are highly bioavailable) and

exposures to early life stages of invertebrates (sperm cells) which often are more sensitive

than adult forms. Tests of sediment pore water toxicity were conducted with the urchin

Arbacia punctulata by the U.S. Geological Survey laboratory in Corpus Christi, Texas.

Sediments from each sampling location were shipped by overnight courier in one-gallon

high density polyethylene jugs chilled in insulated coolers packed with blue ice. Upon

arrival at the laboratory, samples were either refrigerated at 4°C or processed immediately.

All samples were processed (i.e., pore waters extracted) within 10 days of the sampling

date.

Pore water was extracted from sediments with a pressurized squeeze extraction device

(Carr and Chapman, 1992). After extraction, porewater samples were centrifuged in polycarbonate bottles (@1200 g for 20 minutes) to remove any particulate matter. The supernatant was then frozen @-20°C. Two days before the start of a toxicity test, samples were

transferred from a freezer to a refrigerator kept at 4° C, and one day prior to testing, thawed
10

in a tepid (20°C) water bath. Experiments performed by USGS have demonstrated no

effects upon toxicity attributable to freezing and thawing of the pore water samples (Carr

and Chapman, 1992).

Tests followed the methods of Carr and Chapman (1992) and USGS SOP F10.6, developed for Arbacia punctulata. Spawning was induced by a mild electrical shock. Test temperatures of the pore waters, the dilution waters and the tests themselves were maintained

at 20±2° by incubation of in an environmental chamber. Adult A. punctulata were obtained

from Gulf Specimen Co., Panacea, FL. Pore water from sediments collected in Redfish

Bay, Texas, an area located near the testing facility, were used as negative (nontoxic)

controls. Sediment pore waters from this location have been determined to be nontoxic in

this test in repeated trials (Long et al., 1996). Porewater sample salinity was measured and

adjusted to 30±1 ppt, if necessary, using either purified de-ionized water or concentrated

brine. Following these adjustments of salinity, each of the pore water samples was tested

in a dilution series of 100%, 50%, and 25% of the salinity – adjusted sample with 5 replicates per treatment. Dilutions were made with clean, filtered (0.45 um), Port Aransas

laboratory seawater, which has been shown in previous trials to be nontoxic. A dilution

series test with SDS was included as a positive control.

In addition to salinity, other water quality measurements were made for dissolved oxygen,

pH, sulfide and total ammonia. Temperature and dissolved oxygen were measured with

YSI meters; salinity was measured with Reichert or American Optical refractometers; pH,

sulfide and total ammonia (expressed as total ammonia nitrogen, TAN) were measured with

Orion meters and their respective probes. The concentrations of unionized ammonia (UAN)

were calculated using TAN, salinity, temperature, and pH values.

For the sea urchin fertilization test, 50 uL of appropriately diluted sperm were added to

each vial, and incubated at 20±2°C for 30 minutes. One ml of a well mixed dilute egg

suspension was added to each vial, and incubated an additional 30 minutes at 20± 2°C.

Two mls of a 10% solution of buffered formalin solution was added to stop the test. Eggs

with fertilization membranes were counted, and fertilization percentages calculated for each

replicate test.

Microbial bioluminescence (MicrotoxTM) tests. This is a test of the relative toxicity of sediment extracts prepared with an organic solvent. Therefore, it is immune to the effects of

many environmental factors, such as grain size, ammonia and organic carbon. Organic

toxicants, and to a lesser degree trace metals, that may or may not be readily bioavailable

are extracted with the organic solvent. Therefore, this test can be considered as indicative

of the potential toxicity of mixtures of substances bound to the sediment matrices. In previous NS&T Program surveys, the results of Microtox tests have shown extremely high correlations with the concentrations of mixtures of organic compounds. Microtox tests were run

by the U.S. Geological Survey laboratory in Columbia, Mo, on dichloromethane (DCM)

extracts prepared by ABC Laboratories. Final volume adjustments were made with the

carrier solvent dimethylsulfoxide (DMSO). The organic extracts were tested at concentrations of 1.5 to 50 mg equivalent wet wt/mL.
11

The Microtox® assays were performed with dichloromethane (DCM) extracts of sediments

following the basic procedures used in testing sediments from Pensacola Bay and elsewhere (Johnson and Long, 1998). All sediment samples were stored in the dark at 4°C for

5-10 days before processing was initiated. A 3-4 g sediment sample from each station was

weighed, recorded, and placed into a DCM rinsed 50 mL centrifuge tube. A 15 g portion of

sodium sulfate was added to each sample and mixed. Pesticide grade DCM (30 ml) was

added and mixed in. The mixture was shaken for 10 seconds, vented and tumbled overnight.

Sediment samples were allowed to warm to room temperature and the overlying water

discarded. They were then homogenized with a stainless steel spatula, and 15-25 grams

of sediment from each were transferred to a centrifuge tube. The tubes were spun at 1000

g for 5 min. and the pore water was removed using a Pasteur pipette. Three replicate 3-4 g

sediment sub-samples were placed in separate mortars each containing a 15 g portion of

sodium sulfate and mixed. After 30 min sub-samples were ground with a pestle until dry

and added to 50 mL centrifuge tubes. Then, 30 mL of DCM were added to each tube and

shaken to dislodge sediments. Tubes were then shaken overnight on an orbital shaker at a

moderate speed. The tubes were then centrifuged at 500 G for 5 min and the sediment

extracts transferred to Turbovaptm tubes. Then, 20 mL of DCM was added to sediment,

shaken by hand for 10 sec and spun at 500 g for 5 min. The previous step was repeated

and all three extracts were combined in the Turbovaptm tube. Sample extracts were then

placed in the Turbovaptm and reduced to a volume of 0.5 mL. The sides of the Turbovaptm

tubes were then rinsed down with methylene chloride and again reduced to 0.5 mL. Then,

2.5 mL of dimethylsulfoxide (DMSO) were added to the tubes which were returned to the

Turbovaptm for an additional 15 min. Sample extracts were then placed in clean vials and

2.5 mL of DMSO were added to obtain a final volume of 5 mL DMSO.

A suspension of luminescent bacteria, Vibrio fischeri, (Azur Environmental, Inc.) was

thawed and hydrated with toxicant-free distilled water, covered and stored in a 4°C well on

the Microtox analyzer. An aliquot of 10uL of the bacterial suspension was transferred to a

test vial containing the standard diluent (2% NaCl) and equilibrated to 15°C using a temperature-controlled photometer. The amount of light lost per sample was proportional to

the toxicity of that test sample. To determine toxicity, each sample was diluted into four test

concentrations. Percent decrease in luminescence of each cuvette relative to the reagent

blank was calculated. Light loss was expressed as a gamma value and defined as the ratio

of light lost to light remaining.

Because organic sediment extracts were obtained with DCM, a strong nonpolar solvent, the

final extract was evaporated and redissolved in DMSO. DMSO is compatible with the

Microtox system because of its low test toxicity and good solubility with a broad spectrum of

apolar chemicals (Johnson and Long, 1998). The logs of gamma values from these four

dilutions were plotted and compared with the log of the samples’ concentrations. The

concentrations of the extract that inhibited luminescence by 50% after a 5-min exposure

period, the EC50 value, was determined and expressed as mg equivalent sediment wet

weight. Data were reduced using the Microtox Data Reduction software package. All

EC50 values were average 5-min readings with 95% confidence intervals for three replicates.
12

A negative control (extraction blank ) was prepared using DMSO, the test carrier solvent. A

phenol standard (45 mg/L phenol) was run after reconstitution of each vial of freeze-dried V.

fischeri. Tests of extracts of sediments from the Redfish Bay, Texas, site used in the urchin

tests also were used as negative controls in the Microtox tests.

Cytochrome p-450 RGS assays.

All samples were analyzed by the P-450 reporter gene system (RGS) assay, which uses

human liver cells to measure luciferase production in response to activation of CYP1A1

promoter sequences. This assay is responsive to the presence of mixed-function oxidase

inducers such as dioxins, furans, high molecular weight PAHs, and coplanar PCBs in tissues and sediments (Anderson et al., 1995). Therefore, the RGS assay provides an estimate of the presence of contaminants bound to sediment that could produce chronic and/or

carcinogenic effects in benthic biota and/or demersal fishes that feed in sediments. These

tests were run by the Columbia Analytical Services, Inc. in Carlsbad, CA with solvent extracts prepared by their laboratory in Kelso, WA.

 In these tests, standard protocols (Anderson et al., 1996; ASTM, 1997; APHA, 1996) were

followed to ensure comparability with data derived for other areas. Approximately, 20 g of

sediment from each station were extracted by EPA method 3550 to produce 2 mL of

dichloromethane (DCM) extracts. This solvent was exchanged into DMSO, which is less

volatile and less toxic to the test cells. Small portions (5 to 15 uL) were applied to approximately one million human liver cells contained in three replicate wells with 2 mL of culture

medium. After 16 hours of incubation (exposure), the cells were washed, then lysed, and

the solution centrifuged to remove cell debris. Small portions (50 uL) of the cell extracts

were used in measures of luminescence in relative light units (RLU). Solvent blanks and

the reference toxicants (2, 3, 7, 8 – dioxin and benzo[a[pyrene) were tested with each batch

of samples.

Enzyme induction of the standards and samples was calculated (normalized) by dividing

the mean RLU by the mean RLU produced by the solvent blank. The running average fold

induction for dioxin at 3.1 nM (1ng/mL) is approximately 100 and that from 1 ug/mL of

benzo(a)pyrene (b[a]p) is 60 fold. A standard mixture of dioxins and furans, measured as

Toxic Equivalency (TEQ), produces a RGS fold induction equal to the TEQ in pg/mL. Because the primary organic contaminants identified in previous marine studies have been

PAHs, the RGS data were converted to benzo[a]pyrene equivalents (B[a]pEq) by multiplying the fold induction from the applied extract (e.g., 15 uL) by a factor (133.3) to determine

the total in the 2 mL extract, and then dividing by the dry weight of the samples, and 60(=1

ug/gB[a]P). To convert the data to a TEQ(dioxins and furans) in pg/g, the calculations

would be the same, except the factor of 60 would not be used. Tests were conducted with

clean extracts spiked with TCDD and B[a]P to ensure compliance with the results of previous tests.

Chemical analyses – metals.

Chemical analyses of all samples were performed on all 66 samples by the analytical

laboratory at Texas A&M University/Geochemical and Environmental Research Group

(TAMU/GERG) in College Station, Texas on all 66 samples. All analytical methods conformed with performance-based analytical protocols and employed quality-assurance steps
13

of the NS&T Program (Lauenstein and Cantillo, 1993; 1998); including instrument calibration, use of internal standards, replication of some analyses, percent recoveries of spiked

blanks, and analyses of standard reference materials.

Grain size was determined by the standard pipette method following sieving to remove the

sand and gravel fractions. TOC was determined using a Leco Carbon Analyzer. Sediment

samples were digested for final analysis by procedures specific to the instrument method

used. Various concentrating and trapping techniques were used for selected analytes. The

analysis for mercury was performed by cold vapor atomic absorption. Analyses for tin, arsenic, selenium, silver, and cadmium were performed by graphite furnace atomic absorption

spectroscopy. All other metals concentrations were determined by flame atomic absorption

spectroscopy and reported on a dry weight basis. Method detection limits (MDLs) attained in

the analyses are listed in Table 1. Acid-volatile sulfide/simultaneously-extracted metals analyses

were not performed.

Table 1. Trace metals measured in Sabine Lake sediments and method detection

limits (MDLs).

Parameter Method Detection Limit Analytical Method *

(ppm, based on dry weight)

Aluminum 440 FAA

Iron 40 FAA

Manganese 5.0 FAA

Arsenic 0.3 GFAAS

Cadmium 0.008 GFAAS

Chromium 0.1 GFAAS

Copper 0.44 GFAAS

Lead 0.35 GFAAS

Mercury 0.007 CVAA

Nickel 0.7 GFAAS

Selenium 0.2 GFAAS

Silver 0.03 GFAAS

Tin 0.1 GFAAS

Zinc 2.2 FAA

 FAA = Flame atomic absorbtion spectroscopy;

 GFAAS = Graphite furnace atomic absorption spectroscopy

 CVAA = Cold vapor atomic absorption.

Chemical analyses – organic compounds.

The analytes determined in the organic analyses are listed in Table 2, along with their respective MDLs. Sediment samples for organic analysis were prepared by NaSO4 drying, methylene chloride extraction, purified by silica gel/alumina chromatography and concentration.

Quantification was performed using the internal standards method. Polycyclic aromatic hydrocarbons (PAHs) were analyzed by gas chromatography with a mass selective detector in

the selective ion mode. Sediment samples to be analyzed for butyltins were dried with NaSO4

and extracted with methylene chloride containing 2% tropolone, hexylated, purified by silica

gel chromatography, and concentrated. Butyltins were detected by gas chromatography with
14

a tin selective flame photometric detector. Polychlorinated biphenyls and chlorinated pesticides were determined by gas chromatography/electron capture detection. Concentrations

of sediment organic compounds are reported on a dry weight basis.

Table 2. Organic compounds measured in Sabine Lake sediments and method detection limits (MDLs).

Parameter MDL Parameter MDL

(ng/g dry) (ng/g dry)

2,4’Dichloro Diphenyl Ethylene (O,P’DDE) 0.28 Naphthalene 0.5

4,4’Dichloro Diphenyl Ethylene (P,P’DDE) 0.85 C1-Naphthalenes

2,4’Dichloro Diphenyl Dichloroethylene (O,P’DDD) 0.13 C2-Naphthalenes

4,4’Dichloro Diphenyl Dichloroethylene (P,P’DDD) 0.51 C3-Naphthalenes

2,4’Dichloro Diphenyl Trichloroethylene (O,P’DDT) 0.25 C4-Naphthalenes

4,4’Dichloro Diphenyl Trichloroethylene (P,P’DDT) 0.24 1- Methylnaphthalene 0.8

Aldrin 0.25 2- Methylnaphthalene 0.8

Cis-Chlordane 0.66 2,6-Dimethylnaphthalene 2.4

Oxychlordane 2,3,5- Trimethynaphthalene 2.4

Alpha-Chlordane 0.23 Acenaphthalene 3.7

Trans-Nonachlor 0.1 Acenaphthylene 4.5

Cis-Nonachlor Fluorene 2.5

Dieldrin 0.16 C1-Fluorenes

Heptachlor 0.2 C2-Fluorenes

Heptachloro-Epoxide 0.16 C3-Fluorenes

Hexachlorobenzene 0.37 Phenanthrenes 0.5

Alpha-Benzene Hexachloride (HCH) C1-Phenanthrenes

Beta-Benzene Hexachloride (HCH) C2-Phenanthrenes

Lindane (Gamma-Benzene Hexachloride-HCH) 0.22 C3-Phenanthrenes

Delta-Benzene Hexachloride (HCH) 0.17 C4-Phenanthrenes

Endrin 1- Methylphenanthrene 0.6

Mirex 0.08 Anthracene 4.1

Polychlorinated Biphenyls Fluoranthene 0.4

 PCB#8 (CL2) 0.08 Pyrene 3.1

 PCB#18 (CL3) 0.25 Indeno-1,2,3-c,d-Pyrene 1.6

 PCB#28 (CL3) 0.09 Dibenzothiophene

 PCB#44 (CL4) 0.09 C1-Dibenzothiophenes

 PCB#52 (CL4) 0.09 C2-Dibenzothiophenes

 PCB#66 (CL4) 0.14 C3-Dibenzothiophenes

 PCB#101 (CL5) 0.13 C1- Fluoranthene Pyrene

 PCB#105 (CL5) 0.1 Benzo-a-Anthracene 1.4

 PCB#110/77 (CL5/4) * Chrysene 0.5

 PCB#118/108/149 (CL5/5/6) 0.12 C1-Chrysenes

 PCB#128 (CL6) 0.13 C2-Chrysenes

 PCB#138 (CL6) 0.18 C3-Chrysenes

 PCB#126 (CL6) * C4-Chrysenes

 PCB#153 (CL6) 0.12 Benzo-b-Fluoranthene 1.8

 PCB#170 (CL7) 0.81 Benzo-k-Fluoranthene 1.9

 PCB#180 (CL7) 0.16 Benzo-a-Pyrene 1.2

 PCB#187/182/159 (CL7/7/6) 0.14 Benzo-e-Pyrene 2.4

 PCB#195 (CL8) 0.25 Perylene 3.3

 PCB#206 (CL9) 0.09 Benzo-g,h,i-Perylene 0.3

 PCB#209 (CL10) 0.78 Dibenzo-a,h-Anthracene 2.6

Biphenyl 2.4

Chemistry QA/QC.

Quality assurance/quality control (QA/QC) procedures included analyses of duplicates, standard reference materials, and spiked internal standards. In the organic analyses, internal
15

standards were added at the start of the procedure and carried through the extraction, cleanup,

and instrumental analysis steps and used to determine the concentrations of analytes. The

following specific quality assurance steps were used to insure measurement accuracy and

precision:

1. Trace and major metals: Two method blanks and three standard reference materials were

run with each set of no more than 30 samples.

2. Physical/chemical measurements: Grain size duplicates were run every 20 samples. For

TOC, one method blank, one duplicate, and one standard reference material were run

every 20 samples.

3. Pesticides, PCBs and PAHs: One procedural blank, one matrix spike, one duplicate spike

and one standard reference material were run with each batch of no more than 20 samples.

Internal standard recoveries were tracked.

Benthic community.

Benthic community samples were collected at one-third of the station with a petite Ponar

grab sampler. Results of these analyses will be reported in a separate technical report.

Statistical methods.

Several statistical methods were used to identify the significance of the toxicity tests, estimate spatial scales in toxicity and contamination, identify relationships between measures

of toxicity and contamination, and identify chemicals of greatest concern.

Amphipod survival tests. Data from each station in which mean percent survival was less

than that of the control were compared to the Central Long Island Sound control using a

one-way, unpaired t-test (alpha = 0.05) assuming unequal variance. Data were not transformed since examination of data from previous tests have shown that A. abdita percentage survival data meet the requirements for normality.

Significant toxicity for A. abdita is defined here as survival statistically less than that in the

performance control (alpha = 0.05). In addition, samples in which survival was significantly

less than controls and less than 80% of CLIS control values were regarded as “highly

toxic”. The 80% criterion is based upon statistical power curves created from SAIC’s extensive testing database with A. abdita (Thursby et al., 1997) that show that the power to

detect a 20% difference from the control is approximately 90%. The minimum significant

difference (i.e., “MSD” <80% of control response) was used as the critical value in calculations of the spatial extent of toxicity (Long et al., 1996).

Sea urchin fertilization tests. For the sea urchin fertilizations, statistical comparisons

among treatments were made using ANOVA and Dunnett’s one-tailed t-test (which controls

the experiment-wise error rate) on the arcsine square root transformed data with the aid of

SAS (SAS, 1992). The trimmed Spearman-Karber method (Hamilton et al., 1977) with

Abbott’s correction (Morgan, 1992) was used to calculate EC50 (50% effective concentra-
16

tion) values for dilution series tests. Prior to statistical analyses, the transformed data sets

were screened for outliers (Moser and Stevens, 1992). Outliers were detected by comparing the studentized residuals to a critical value from a t-distribution chosen using a

Bonferroni-type adjustment. The adjustment is based on the number of observations (n) so

that the overall probability of a type 1 error is at most 5%. The critical value (CV) is given

by the following equation: cv= t(dfError, .05/[2 x n]). After omitting outliers but prior to

further analyses, the transformed data sets were tested for normality and for homogeneity

of variance using SAS/LAB Software (SAS, 1992). Statistical comparisons were made with

mean results from the Redfish Bay controls. Reference toxicant concentration results were

compared to filtered seawater controls and each other using both Dunnett’s t-test and

Duncan’s multiple range test to determine lowest observable effects concentrations

(LOECs) and no observable effects concentrations (NOECs).

In addition to the Dunnett’s one-tailed t-tests, data from field-collected samples were

treated with an analysis similar to the MSD analysis used in the amphipod tests. Power

analyses of the sea urchin fertilization data have shown MSDs of 15.5% for alpha = 0.05

and 19% for alpha = 0.01. However, to be consistent with the statistical methods used in

previous surveys (Long et al., 1996), estimates of the spatial extent of toxicity were based

upon the same critical value used in the amphipod tests (i.e., <80% of control response).

Microtox tests. Microtox data were analyzed using the computer software package developed by Microbics Corporation to determine concentrations of the extract that inhibit luminescence by 50% (EC50). This value was then converted to mg dry wt. using the calculated dry weight of sediment present in the original extract. To determine significant differences of samples from each station, pair-wise comparisons were made between survey

samples and results from Redfish Bay control sediments using analysis of variance

(ANOVA). Concentrations tested were expressed as mg dry wt based on the percentage

extract in the 1 ml exposure volume and the calculated dry wt of the extracted sediment.

Statistical comparisons among treatments were made using ANOVA and Dunnett’s onetailed t-tests on the log transformed data with the aid of SAS (SAS, 1992).

Two additional critical values were generated for the Microtox test results, both based upon

statistical analyses of the existing data from NOAA surveys conducted thus far (n=1013).

The two critical values are <0.06 mg/ml and <0.51 mg/mL. The first value (0.06 mg/mL)

represented the 90% lower prediction limit (LPL) of the entire data set. The probability that

a future observation from this data distribution would be less toxic (i.e., have a greater

EC50 value) than 0.06 mg/ml would be 90%. Therefore, a sample with an EC50 less than

0.06 mg/mL would be considered to be extremely toxic in this test, but should occur rarely.

The second value (0.51 mg/mL) represents the 80% LPL following removal of the lowest

(most toxic) 10% of the data values from the database to eliminate the affects of these

extremely toxic samples upon the distribution of the data.

Cytochrome P-450 RGS tests. Microsoft Excel 5.0 was used to determine the mean RGS

response and the 99% confidence interval of the b[a]p equivalent values for all 66 samples.

Two critical values were calculated and used to estimate spatial extent of toxicity in this

test. The first value, 37.1 ug/g benzo(a)pyrene equivalents, represented the upper 90%

prediction limit (UPL) of the entire data set gathered thus far in all NOAA studies (n=530).
17

This value agrees well with 32 ug/g, the RGS induction level equivalent to the ERL value

(Long et al., 1995) for high molecular weight PAHs determined in regression analyses of

the existing data for this test. Therefore, this value is viewed as a concentration above

which toxicologically significant effects may begin in sediments. The second value, 11.1

ug/g, was the 80% UPL of the data distribution following elimination of the data above the

90th percentile of the entire data base. This value (11.1 ug/g) is viewed as the upper limit of

background RGS responses.

Spatial patterns in toxicity Spatial patterns in toxicity were illustrated by plotting toxicity

data on base maps of each major region. The incidence of toxicity was determined by

dividing the numbers of samples identified as significantly different from controls or “highly

toxic” by the total numbers of samples (n=66) tested.

Estimates of the spatial extent of toxicity were determined with cumulative distribution

functions in which the toxicity results from each station were weighted to the dimensions

(km2) of the sampling stratum in which the samples were collected (Schimmel et al., 1994).

The size of each stratum (km2) was determined by use of an electronic planimeter applied

to navigation charts, upon which the boundaries of each stratum were outlined. Stratum

sizes were calculated as the averages of three trials, all of which were within 10% of each

other. A critical value of less than 80% of control response was used in the calculations of

the spatial extent of toxicity for all tests. That is, the sample-weighted sizes of each stratum

in which toxicity test results were less than 80% of control responses were summed to

estimate the spatial extent of toxicity. In addition, the critical values described above for the

Microtox and RGS assays were used in these analyses.

Spatial patterns in chemical concentrations. Chemical data from the sample analyses

were plotted on base maps to identify spatial patterns in concentrations. Maps were prepared for several chemicals representative of the different chemical classes. The spatial

extent of contamination was determined with cumulative distribution functions in which the

sizes of strata in which samples exceeded effects-based, numerical guidelines were

summed, using the Effects Range-Median (ERM) values of Long et al. (1995).

Chemistry/toxicity relationships. Chemistry/toxicity relationships were determined in a

multistep sequence. First, nonparametric, Spearman-rank correlations were determined

(Statview software) for each toxicity test and each physical/chemical variable. The correlation coefficients (rho) and their statistical significance (p values) were recorded and compared among chemicals to identify which chemicals co-varied with toxicity and which did

not.

Second, for those chemicals in which a significant correlation was observed, the data were

examined in scatterplots to determine if there was a reasonable pattern of increasing toxicity with increasing chemical concentration and if any chemical in the toxic samples equalled

or exceeded published numerical guidelines. In this step, chemical concentrations in the

scatterplots were compared with the ERM values to determine which samples, if any, were

both toxic and had elevated chemical concentrations. The concentrations of unionized

ammonia were compared to lowest observable effects concentrations (LOEC) determined

for the sea urchin tests by the USGS (Carr et al., 1995) and no observable effects concentrations (NOEC) determined for amphipod survival tests (Kohn et al.,1994).
18

Third, the numbers (and percentages) of samples out of those that were analyzed that

exceeded the respective guidelines were determined. The combined results of these steps

were examined to determine which chemical(s), if any, may have contributed to the observed toxicity and which probably had a minor or no role in toxicity.

Correlations were determined for all the substances that were quantified, including total

(bulk) trace metals, metalloids, unionized ammonia (UAN), percent fines, total organic

carbon (TOC), chlorinated organic hydrocarbons (COHs), and polynuclear aromatic hydrocarbons (PAHs). In addition, a chemical index calculated as the sums of quotients formed

by dividing the chemical concentrations in the samples by their respective ERM values

were compared to the results of the toxicity tests. Those substances that showed significant correlations were indicated with asterisks (*for p<0.05, ** for p<0.01, ***for p<0.001,

and ****for p<0.0001).

In correlation analyses involving a large number of variables such as in this survey, some

correlations could appear to be significant by random chance alone. Statistical adjustments

often are needed to account for this possibility. Accordingly, readers should note that in the

results tables only those correlations shown with either three or four asterisks would remain

significant if the number of variables (50) were taken into account in these analyses. However, the correlation coefficients are shown without these adjustments.

RESULTS

Severity and spatial patterns in toxicity.

Results of the amphipod, sea urchin, Microtox, and cytochrome P-450 RGS tests are summarized for each sampling station in Table 3. Except in the case of the RGS assay, results

are expressed as percentages of negative controls. RGS assay results are expressed as

ug benzo(a)pyrene equivalents/g sediment.

Mean amphipod survival among the 66 samples ranged from 85% to 110% of negative

controls. Amphipod survival was reduced significantly relative to controls in 13 samples;

however, mean survival exceeded 80% of controls in all samples. Microtox test results

were significantly different from controls in 52 samples and sample means ranged from

1.4% to 210% of controls. EC50 values were less than 0.51 mg/mL (indicating moderate

loss in metabolic activity) in three samples. None of the mean EC50 values were less than

0.06 mg/mL, which would have indicated a severe response. In the sea urchin tests, mean

percent fertilization was significantly reduced in 15 samples, 6 samples, and none of the

samples in tests performed with 100%, 50%, and 25% pore water concentrations, respectively. In tests of 100% pore water, percent fertilization ranged from 4% to 108% of controls. Percent normal development was significantly reduced in 62 samples, 45 samples,

and one of the samples in tests of 100%, 50%, and 25% pore water concentrations, respectively. Cytochrome P-450 RGS assay results ranged from 1.0 to 104 ug/g; responses

exceeded 11.1 ug/g in 21 samples (indicating moderate induction) and exceeded 37.1 ug/g

in 6 samples (indicating high induction).

The 13 samples in which mean amphipod survival was significantly less than controls are

illustrated in Figure 4. Results were significant in only three stations within the lake
19

Table 3. Summary of toxicity test results for 66 sediment sampling stations in Sabine Lake.

Stratum Sta.

Ampelisca abdita 

%survival Microtox (EC50)

Arbacia punctulata 

%fertilization (100%)

Arbacia punctulata 

%fertilization (50%)

Arbacia punctulata 

%fertilization (25%)

Arbacia punctulata 

%normal development 

(100%)

Mean % of ctrl Sign. Mean % of ctrl Sign. Mean % of ctrl Sign. Mean % of ctrl Sign. Mean % of ctrl Sign. Mean % of ctrl Sign. Mean % of ctrl Sign. Mean % of ctrl Sign.

A 1 88 96 0.45 1.46 S 94.2 108 93.8 104 94.8 110 0 0 S 1 1 S 91.2 100 6.21

A 2 93 101 0.71 2.31 S 72.8 83 S 91.2 101 92.6 107 0 0 S 0 0 S 93.8 103 61.10

A 3 88 96 0.68 2.21 S 81.6 93 97.6 108 91 105 0 0 S 0 0 S 80.8 89 63.45

B 4 90 98 0.9 2.93 S 89.6 102 94.6 105 88 102 0 0 S 89 95 91.8 101 15.20

B 5 83 90 5.33 17.34 S 84.6 97 84.6 94 78 90 89 101 90.8 97 90.4 99 6.51

B 6 88 96 34.07 110.87 91.2 104 91.4 101 95.8 111 0 0 S 0 0 S 89.8 99 12.44

C 7 87 94 7.67 24.96 S 89.4 102 86 95 79.6 92 84.2 96 87.6 94 92.4 102 20.85

C 8 83 89 S 23.07 75.07 85.6 98 87.4 97 86.4 100 0 0 S 0 0 S 92.6 102 20.92

C 9 85 91 S 10.7 34.82 S 91 104 90.2 100 86.2 100 0 0 S 79.2 85 89.2 98 28.18

D 10 91 99 0.48 1.56 S 74.6 85 S 72.8 81 S 80.3 93 0 0 S 0 0 S 91.6 101 16.08

D 11 87 95 2.83 9.21 S 86.8 99 85.8 95 83.5 97 1.8 2 S 88.8 95 86.2 95 53.68

D 12 81 88 S 0.57 1.85 S 36.8 42 S 91 101 88 102 0.2 0 S 0 0 S 87.6 96 71.67

E 13 90 97 13.43 43.7 S 84.4 96 90.2 100 91.4 106 88.2 100 91.3 98 90.4 99 2.20

E 14 79 85 S 45.93 149.46 68.4 78 S 70.2 78 S 82.8 96 81.6 93 82 88 86 95 7.25

E 15 84 90 S 6.27 20.4 S 91.8 105 94.6 105 90.2 104 0.4 0 S 3 3 S 86.8 95 14.20

F 16 90 97 22.6 73.54 94.8 108 93.8 104 93.8 109 0 0 S 0.2 0 S 87.8 96 10.69

F 17 91 98 5.17 16.82 S 89.2 102 89.6 99 92.6 107 0 0 S 0 0 S 85.2 94 13.70

F 18 88 95 4.4 14.32 S 94.4 108 94.5 105 93.6 108 0.4 0 S 17.8 19 S 88 97 9.86

G 19 89 96 15.5 50.44 S 91.8 105 91.6 102 91.8 106 0 0 S 0 0 S 93 102 28.21

G 20 88 95 S 15.97 51.97 S 90.2 103 91.8 102 91.8 106 0.4 0 S 0.5 1 S 90.8 100 22.86

G 21 88 95 17.13 55.74 S 92.6 106 94.4 105 92 106 0.4 0 S 92.4 99 93.4 103 33.07

H 22 87 94 42.43 138.07 81 92 88.2 98 85.2 99 0 0 S 83 89 91.8 101 8.05

H 23 81 87 S 22.63 73.64 84.2 96 88.8 99 88.2 102 0.8 1 S 85.8 92 91.6 101 9.43

H 24 82 88 S 6.18 20.11 S 94 107 91.2 101 89.4 103 0 0 S 2.8 3 S 88.6 97 13.90

I 25 89 96 31.73 103.25 88 100 90.8 101 89.8 104 0 0 S 0 0 S 84 92 13.85

I 26 95 102 5.77 18.78 S 90.8 104 91 101 88.6 103 0 0 S 34 36 S 86 95 14.16

I 27 81 89 S 64.73 210.64 85.4 97 93 103 88.8 103 0 0 S 14 15 S 89.2 98 6.63

J 28 77 85 0.85 2.77 S 36.6 49 S 63.8 88 69.4 102 0 0 S 39.8 41 S 91.8 95 91.07

J 29 85 93 2.03 6.61 S 43 58 S 65.6 90 73.4 108 0 0 S 20.8 22 S 84 87 103.76

J 30 81 89 S 3.57 11.62 S 67.6 91 70.2 97 68.8 101 0 0 S 0 0 S 87.6 91 18.78

K 31 86 95 1.39 4.56 S 73.6 84 90.8 101 85.6 99 0 0 S 0 0 S 87 96 9.70

K 32 81 89 38.3 124.63 89 102 90.8 101 90.4 105 0 0 S 84.6 91 81.2 89 4.63

P450 BaP

EQ(ug/g)

Arbacia punctulata

% normal development

(50%)

Arbacia punctulata

% normal development

(25%)
20

Table 3. Summary of toxicity test results for 66 sediment sampling stations in Sabine Lake continued.

Stratum Sta.

Ampelisca abdita 

%survival Microtox (EC50)

Arbacia punctulata 

%fertilization (100%)

Arbacia punctulata 

%fertilization (50%)

Arbacia punctulata 

%fertilization (25%)

Arbacia punctulata 

%normal development 

(100%)

Mean % of ctrl Sign. Mean % of ctrl Sign. Mean % of ctrl Sign. Mean % of ctrl Sign. Mean % of ctrl Sign. Mean % of ctrl Sign. Mean % of ctrl Sign. Mean % of ctrl Sign.

M1 37 95 98 6.47 21.05 S 69.2 93 74.4 102 71.6 105 62.4 65 S 84.4 88 92.8 96 3.19

M1 38 89 92 S 6.03 19.62 S 61.6 83 62.2 86 71.6 105 58 60 S 77 80 S 94.8 98 6.57

M1 39 96 99 5.77 18.78 S 70 94 66 91 60.4 89 78.8 82 S 89.8 93 97.4 101 2.86

M2 40 96 99 3.67 11.94 S 69.8 94 64 88 66.4 98 59.8 62 S 79.4 82 S 95.6 99 10.98

M2 41 97 100 24.87 80.93 76.4 103 71 98 67.8 100 57.6 60 S 85.2 88 94.6 98 5.84

M2 42 93 96 10.53 34.27 S 79.2 107 75.8 104 73 108 58.8 61 S 83.8 87 93.4 97 5.58

M3 43 95 98 19.67 64.01 S 73.8 100 72.8 100 72.8 107 46.6 49 S 75.4 78 S 93.6 97 7.21

M3 44 98 101 32.53 105.86 73.8 100 70.8 98 65.8 97 33.2 35 S 67.6 70 S 89.4 93 7.45

M3 45 96 99 49.4 160.75 76.4 103 78.8 109 66.2 97 56 58 S 81.8 85 93.6 97 5.23

M4 46 88 96 9.43 30.69 S 74.6 101 68.4 94 72.2 106 38.6 40 S 72.4 75 S 88.4 92 1.84

M4 47 86 93 6.2 20.18 S 62.4 84 63 87 67.6 100 34.8 36 S 67.8 70 S 87 90 2.20

M4 48 79 86 S 9.4 30.59 S 69.6 94 75.4 104 76.4 113 43 45 S 77.4 80 S 92.2 96 2.12

M5 49 87 95 22.17 72.14 80 108 78.6 108 74.2 109 8.3 9 S 65.4 68 S 86.6 90 2.11

M5 50 93 101 8.6 27.99 S 68 92 67.8 93 72.8 107 37 39 S 68 71 S 89 92 2.07

M5 51 86 93 10.13 32.96 S 66 89 71.2 98 72.8 107 44.4 46 S 76.8 80 S 93.8 97 1.99

M6 52 83 90 5.3 17.25 S 65.8 89 71 98 65.6 97 60.8 63 S 83 86 95 99 3.10

M6 53 86 93 9.7 31.57 S 65.6 89 67.2 93 68.4 101 49.6 52 S 72.2 75 S 92.4 96 3.89

M6 54 88 96 6.77 22.03 S 63.4 86 71.2 98 67.4 99 68.6 72 S 87.8 91 96.2 100 6.32

M7 55 78 85 3 9.76 S 67.2 91 67.4 93 74.6 110 68 71 S 91 94 96.6 100 2.13

M7 56 81 88 S 3.93 12.79 S 71.6 97 69.4 96 73.4 108 73 76 S 90.4 94 97.6 101 4.88

M7 57 88 96 49.03 159.55 66.4 90 72.6 100 75.6 111 60 63 S 87.6 91 95.6 99 1.46

O 58 92 97 9.8 31.89 S 50.8 69 S 66.6 92 69.6 103 0 0 S 56.4 59 S 84.8 88 3.12

O 59 85 104 3.47 11.29 S 47.8 65 S 55.8 77 S 58.2 86 0.2 0 S 69.8 72 S 85.4 89 2.51

O 60 90 110 1.47 4.78 S 54.4 73 S 64.4 89 74 109 0 0 S 57.4 60 S 85.6 89 2.75

P 61 95 98 0.66 2.15 S 47.2 64 S 72 99 73.8 109 0 0 S 62.8 65 S 86.4 90 2.28

P 62 98 101 1.3 4.23 S 61.8 83 S 57.4 79 64.4 95 0 0 S 65 67 S 87.8 91 5.15

P 63 99 102 1.97 6.41 S 70 94 71 98 61.8 91 0 0 S 58 60 S 77 80 S 3.69

R 64 94 99 0.76 2.47 S 3.2 4 S 49.6 68 60.8 90 0 0 S 76.6 80 S 95.2 99 1.02

R 65 89 109 0.44 1.43 S 38.4 52 S 65.2 90 75 110 0 0 S 80.6 84 S 92 96 

R 66 86 104 1.18 3.84 S 49.6 67 S 58.4 80 S 58.8 87 0 0 S 66.4 69 S 84.4 88 3.08

REF 01 87.6 100 R 90.1 100 R 86.4 100 R 87.8 100 R 93.3 100 R 91 100 R

REF 02 74.1 100 R 72.6 100 R 67.9 100 R 95.9 100 R 96.3 100 R 96.3 100 R

REF 03 30.73 100 R

REF 04 82 0 R

REF 05 91 0 R

REF 06 92 0 R

REF 07 93 0 R

REF 08 95 0 R

REF 09 97 0 R

P450 BaP

EQ(ug/g)

Arbacia punctulata

% normal development

(50%)

Arbacia punctulata

% normal development

(25%)
21

1

2

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7 8 9

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56

61

amphipod survival not significantly

different from controls (p>0.05)

amphipod survival significantly different

from controls

amphipod survival significantly different

from controls and <80% of controls

Figure 4. Sampling stations in which results of amphipod survival tests were not

significant, significant only, or highly significant.
22

whereas survival was significantly reduced in 10 samples from the channels of the Intracoastal Waterway (ICW). Significantly reduced survival occurred in five of the samples from

the Sabine-Neches Canal (stations 19 through 33). None of the stations in the lower portions of the ICW and entrance channel were significantly toxic in this test. Based upon the

statistically determined criterion of <80% of control survival (Thursby et al., 1997), none of

the samples was highly toxic in the amphipod tests.

As observed in the amphipod tests, toxicity in the urchin fertilization tests was scattered

among the stations (Figure 5). None of the samples collected in Sabine Lake were toxic in

the urchin fertilization tests. Several samples from the upper reaches of the Neches River

(stations 1 – 9) were toxic in 100% pore water. One sample each from the Neches and

Sabine rivers was toxic in both 100% and 50% pore water concentrations. Toxicity was

much more apparent in samples from the Sabine Pass and entrance channel area (stations

34 - 35, and 58 – 66). However, none of the samples was toxic in 25% pore water concentrations.

Toxicity was much more pervasive and severe as measured in the urchin embryo development tests (Figure 6) as compared to the other tests done with invertebrates. A large

majority (45 of 66) of the samples were toxic in at least the 100% and 50% pore water

concentrations. All samples seaward of the confluence of the Sabine-Neches canal and

the Sabine Lake channel were toxic. The only sample in which a significant response was

observed in 25% pore water was from an offshore station (number 63).

Microtox data are illustrated as histograms in Figure 7 to more clearly identify the spatial

patterns in response. The data are shown as the ratio of the sample mean to the reference

(Redfish Bay) mean; therefore, the toxic responses increased with the height of the bars.

For example, the ratio between the sample mean for station 1 (EC50 = 0.45 mg/mL) and

the reference mean (EC50 = 31 mg/mL) was 68, indicating a considerably smaller amount

of extract caused a 50% reduction in bioluminescence. In contrast, the EC50 for station 6

was 34 mg/mL, resulting in a ratio to reference mean of 0.9. The data indicated there were

three areas in which bioluminescence activity was reduced the most. These areas were

the upper Neches River (stations 1- 4 and 10 – 12), the Taylor Basin and vicinity (stations

28 – 31), and the offshore area (stations 60 – 66). Sample means were less than 0.51 mg/

ml at three stations (1, 10, and 65), indicating a moderate degree of response. Samples

collected in Sabine Lake were among the least toxic.

Results of the RGS assays indicated that stations from the upper Neches River and Taylor

Basin gave the highest induction responses (Figure 8); however, unlike the Microtox tests,

the samples from the offshore stations did not cause relatively high responses. RGS assay

results exceeded 37.1 ug/g in samples from four stations in the Neches River and two

stations in Taylor Basin, indicating a relatively high level of response. As observed in the

Microtox tests, samples collected in Sabine Lake were among the least toxic.

Spatial extent of toxicity.

The spatial extent of toxicity was estimated by weighting the data from the toxicity tests to

the sizes of the sampling strata and calculating a sum of the strata sizes in which toxic

responses were observed. Critical values (or criteria) for defining samples as “toxic” are

listed for each test. The estimates are summarized in Table 4.
23

1

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4

5

6

7 8 9

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30

40

35

36

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59

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49

57

14

52

33

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25

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29

60

15

26

34

44

53

63

27

64

65

56

61

Figure 5. Sampling stations in which results of urchin fertilization tests were either not 

significant, or significant in 100%, 50%, or 25% pore water concentrations.

urchin fertilization not significant

 in 100% pore water

urchin fertilization significant

 in only 100% pore water

urchin fertilization significant

 in 100% + 50% pore water

urchin fertilization significant

 in all porewater concentrations
24

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25

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29

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15

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34

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53

63

27

64

65

56

61

Figure 6. Sampling stations in which results of urchin development tests were either not 

significant, or significant (p<0.05) in 100%, 50%, or 25% pore water concentrations.

urchin development not significant

 in 100% pore water

urchin development significant

 in only 100% pore water

urchin development significant

 in 100% + 50% pore water

urchin development significant

 in 100% + 50% + 25% pore water 
25

1

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3 4

5 6

7 8 9

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53

63

27

64

65

56

61

Figure 7. Results of Microtox tests expressed as ratio of sample mean to reference

mean.

0

10

20

30

40

50

60

70

80

Ratio of

sample

mean to

reference mean
26

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18 17

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54 29

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64

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56

61

Figure 8. Results of cytochrome P-450 RGS assays expressed as benzo(a)pyrene

equivalents.

0

20

40

60

80

100

120

Benzo(a)

pyrene

equivalents,

ug/g dry wt.
27

Table 4. Estimates of the spatial extent of toxicity in sediments of Sabine 

Lake based upon results from five independent tests.

Toxic area Percent 

Toxicity test Criterion (Km2

) of total*

Percent amphipod survival

• <80% of control 0.0 0.0

Percent urchin fertilization

• <80% of control

in 100% pore water 14.5 5.9

• <80% of control

in 50% pore water 7.3 2.9

• <80% of control

in 25% pore water 0.0 0.0

Percent normal urchin development

• <80% of control

in 100% pore water 244.6A 99.4A

• <80% of control

in 50% pore water 107.6B 43.4B

• <80% of control

in 25% pore water 0.0 0.0

Microtox bioluminescence EC50

• <80% of control 194.2 78.9

• <0.51 mg/ml 3.6 1.4

• <0.06 mg/ml 0.0 0.0

Cytochrome p-450 induction

• >11.1 ug/g 6.7C 2.7C

• >37.1 ug/g 1.7D 0.7D

* total area: 246 km2

A Toxic area: 243.4 km2

 (99.4% of 244.8 km2) accounting for second

alternates at two stations B Toxic area: 107.4 km2

 (43.7% of 245.8 km2) accounting for second

alternate at one station

C Toxic area: 6.5 km2

 (2.6% of 245.8 km2) accounting for second

alternate at one station D Toxic area: 1.5 km2

 (0.6% of 245.8 km2) accounting for second

alternate at one station

Amphipod survival exceeded 80% of controls in all samples; therefore, the spatial extent of

toxicity was estimated as 0% in that particular test. In the sea urchin tests of fertilization

success, strata in which percent fertilization was less than 80% of controls represented 6%,

3%, and 0% of the area in tests of 100%, 50%, and 25% pore waters, respectively. In the

urchin embryo development tests, results were less than 80% of controls in strata that

represented 99%, 43%, and 0% of the area in the three pore water concentrations. Similarly, because of the large differences between the Redfish Bay control means and sample

means in the Microtox tests, the spatial extent of toxicity as defined as <80% of controls

was 79%. However, when defined as a EC50 response of <0.51 mg/mL and <0.06 mg/mL,
28

the estimates of the spatial extent of toxicity were 1.4% and 0%, respectively. Strata in

which cytochrome p-450 RGS responses exceeded 11.1 ug/g and 37.1 ug/g represented

3% and 1% of the total area, respectively.

There were two sampling stations in which the first alternate set of coordinates were not

sampled. At station 3, an active dredging operation precluded access to the target location.

At station 57, oyster shells pinned the jaws of the sampler open and precluded collection of

sediments. Therefore, samples were collected at second alternate locations for both stations. Estimates of the spatial extent of toxicity were adjusted for this situation by re-calibrating the areas of the strata in which these two stations were located (Table 4). The area

of stratum A-3 was reduced by 0.23 km2 (0.46 km2 divided by 2) and the area of stratum M7 was reduced by 1.0 km2

 (4.0 km2 /4). The estimates based upon the adjusted areas are

shown as footnotes on Table 4 for the tests in which the results were affected.

The estimates of spatial extent of toxicity in tests other than those with amphipods were

highly influenced by the data from the Redfish Bay controls which proved to be unusually

non-toxic. This situation was observed in analyses of data from another NOAA survey

conducted in Puget Sound. Thus, the comparisons with the control responses probably

exaggerated the degree of toxicity. Estimates based upon numerical critical values derived

from analyses of the distribution of results in these tests in previous NOAA surveys are

probably more reliable.

Summary of toxicity results. The results of the toxicity tests indicated that sediments in

this survey area were not highly toxic (i.e., percent survival > 80% of controls) as measured

with the acute amphipod survival tests. However, amphipod survival was significantly

reduced in 13 samples and results of the sublethal tests performed with the urchin gametes

and embryos were significant in additional samples. Urchin embryo development was

significantly reduced in most samples. These data, collectively, suggest that sediments

were slightly to moderately toxic, but not highly toxic, in parts of the study area. Results of

the two tests performed with organic solvent extracts of the sediments showed that the

potential for toxicity was highest in upper Neches River, Taylor Basin, and regions offshore

beyond the entrance channel. Highly significant responses in these two tests, however,

were neither pervasive nor widespread.

Spatial patterns in chemical concentrations. Distinct spatial patterns in chemical concentrations were difficult to identify because of the similarity in the data among the 66

stations. Most substances indicated very little variation in concentrations throughout the

study area. For example, the concentrations of the trace metal arsenic ranged from 0.3 to

15 ppm, considerably below the ERM value of 70 ppm. Arsenic concentrations were higher

in the Sabine-Neches canal (stations 6-27) than in the adjacent basin of Sabine Lake

(Figure 9). Arsenic concentrations also were relatively high offshore (stations 58-66) as

compared to the Sabine Lake stations. Although arsenic concentrations exceeded the ERL

value of 8.2 ppm in many samples; none equalled or exceeded the ERM value of 70 ppm.

The pattern in the concentrations of total PAHs (sum of 13 compounds for which numerical

guidelines have been derived) was clearer than that for arsenic. Total PAH concentrations

were highest in samples from stations 2, 11, and 12 in the upper Neches River and from

station 28 – 30 in the Taylor Basin (Figure 10). As observed with arsenic, total PAH con-
29

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Figure 9. Concentrations of arsenic in sediments from Sabine Lake and vicinity.

0

3

6

9

12

15

ERL=

8.2

ppm
30

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Figure 10. Concentrations of total PAHs in sediments from Sabine Lake and vicinity.

0

1000

2000

3000

4000

5000

6000

total PAHs, 

ppb

ERL = 

4022
31

centrations were somewhat higher in the canals than in the adjacent basin of Sabine Lake.

However, unlike the pattern with arsenic, PAH concentrations decreased seaward and

offshore.

Spatial extent of chemical contamination. The spatial extent of chemical contamination

was estimated for 25 substances and/or classes of compounds by weighting the data to the

sizes of the sampling strata, similar to the methods used to calculate the spatial extent of

toxicity. ERL and ERM values from Long et al. (1995) were used as the critical values

(Table 5). The data indicated that only one substance, acenaphthalene, equalled or exceeded an ERM value. The one sample in which the ERM value was exceeded represented 0.3 km2

, equivalent to 0.1% of the total study area. Concentrations of arsenic and

nickel were elevated above the respective ERL values in the most stations, representing

about 17% and 9% of the study area, respectively. The spatial extent of contamination by

all other substances was less than 1.0% of the area. Similarly, the samples in which the

mean ERM quotients exceeded 0.1 represented less than 1% of the area (Table 5).

Relationships between toxicity and chemical concentrations. The relationships between results of laboratory toxicity tests and concentrations of chemical substances in the

sediments were determined with correlation analyses. Initially, correlations were calculated

for classes of chemicals normalized to their respective ERM values and expressed as

mean chemical:ERM quotients (Table 6) to identify which groups of chemicals, if any, covaried with measures of toxicity. Chemical classes were designated as nine trace metals,

thirteen PAHs, three chlorinated organic hydrocarbons, and all 25 substances combined.

None of the chemical classes showed significant relationships with either the amphipod

survival or urchin fertilization tests. Microtox results were significantly correlated, but only

at a significant level of 0.05, with the mean ERM quotients for all 25 substances.

In contrast, the concentrations of all chemical groups were highly correlated with the urchin

development and P-450 RGS test results. In the urchin development tests, the strongest

correlations were with the trace metals. In the P-450 RGS assays, the strongest relationships were with the PAHs and to a lesser degree the chlorinated organics. These correlations do not establish a causative relationship, but, rather, a correlative association between a set of independent variables (chemical concentrations) and dependent variables

(toxicity tests).

To further identify which organic compounds showed strong relationships with the results of

the cytochrome P-450 RGS assays, correlation analyses were continued for a variety of

different chemical classes (Table 7). Correlations were highly significant for all classes of

chemicals: PAHs, DDTs, HCHs, and PCBs; therefore, indicating that results of this test covaried with mixtures of several classes of organics in the sediments.

To further elucidate the associations between urchin development and trace metals, correlations were performed for individual trace metals (Table 8). The results also include the

correlation coefficient for unionized ammonia, a substance commonly found in sediments

that can be toxic. The correlations of percent normal development with concentrations of

ammonia and all trace metals were significant. They were significant, except for silver, at a

probability level of 0.0001 or greater. Zinc and tin showed the strongest relationships with

percent normal development (i.e., Spearman-rank, rho>0.7).
32

Table 6. Spearman-rank correlation coefficients (rho, corrected for ties) for toxicity test results 

and chemical concentrations normalized to ERM values in Sabine Lake sediments (n=66).

Percent Percent urchin Percent urchin Cytochrome

Chemical amphipod fertilization normal development Microtox P-450

group survival (100% pore water) (100% pore water) EC50 RGS

mean ERM quotient: metals 0.012 -0.119 -0.698 **** -0.19 -0.429 * * *

mean ERM quotient: PAHs -0.173 0.168 -0.408 * * * -0.13 0.827 ***

mean ERM quotient: COHs 0.080 0.126 -0.527   -0.13 0.779 ** *

mean ERM quotient: all -0.040 -0.094 -0.677 **** -0.28 * 0.662 ***

* p<0.05

** p<0.01

***p<0.001

****p<0.0001

Table 5. Estimates of the spatial extent of chemical contamination relative to

effects - based numerical guidelines (ERL and ERM) for Sabine Lake. Data

are shown as km2

 and percentages of total study area*.

Chemical >ERL >ERM

or chemical class km2 Pct. of total km2 Pct. of total

arsenic 40.7 16.5 0 0

cadmium 0 0 0 0

chromium 0 0 0 0

copper 0 0 0 0

lead 0 0 0 0

mercury 0 0 0 0

nickel 22.2 9 0 0

silver 0 0 0 0

zinc 0 0 0 0

naphthalene 0 0 0 0

2-methyl naphthalene 0.2 0.1 0 0

acenaphthylene 0 0 0 0

acenaphthene 1.7 0.7 0.3 0.1

fluorene 1.7 0.7 0 0

phenanthrene 0.3 0.1 0 0

anthracene 0.9 0.4 0 0

fluoranthene 0.9 0.4 0 0

pyrene 0.5 0.2 0 0

benzo(a)anthracene 1.1 0.4 0 0

chrysene 0 0 0 0

benzo(a)pyrene 0 0 0 0

dibenz(a,h)anthracene 0.4 0.2 0 0

sum low PAHs 0.9 0.4 0 0

sum high PAHs 0.9 0.4 0 0

total PAHs 0.5 0.2 0 0

p,p' - DDE 0 0 0 0

total DDTs 1 0.4 0 0

total PCBs 0.8 0.3 0 0

mean ERM quotient >0.1 0.1 0.4

total area: 246 km2
33

Table 7. Spearman-rank correlation coefficients (rho, corrected for ties) 

for P-450 RGS test results and concentrations

of classes of organic compounds in Sabine Lake sediments (n=66).

Class of P-450 RGS

compounds induction

sum 7 LPAHs 0.802 ****

total LPAHs 0.794  

sum 6 HPAHs 0.828 ** **

total HPAHs 0.812  

total 13 PAHs 0.827 ** **

total PAHs 0.814  

total DDTs 0.840 ** **

total HCHs 0.583  

total PCBs 0.779 ** **

 p<0.0001

Table 8. Spearman-rank correlation coefficients (rho,corrected for ties)

for percent urchin normal development (100% porewater) and concentrations

 of trace metals and un-ionized ammonia in Sabine Lake sediments (n=66).

Chemical Percent sea urchin

name normal development

un-ionized ammonia -0.547 ** **

arsenic -0.665  

cadmium -0.583 ** **

chromium -0.521  

copper -0.739 ** **

iron -0.714  

lead -0.632 ** **

manganese -0.646  

mercury -0.600 ** **

nickel -0.655  

selenium -0.592 ** **

silver -0.270 *

tin -0.731 ****

zinc -0.699 ****

*p<0.05

**p<0.01

***p<0.001

****p<0.0001
34

The Lowest Observable Effects Concentrations (LOEC) for unionized ammonia in the

fertilization test (800 ug/L) was not exceeded in the Sabine Lake samples. However, the

LOEC of 90 ug/L for the embryological development test was exceeded in 35 of the 66

samples. Therefore, the data suggest that ammonia alone may have significantly contributed to the toxicity observed in many of the samples tested for embryological development.

Summary of chemical contamination. The chemical data for a wide variety of chemical

substances indicated that the surficial sediments in Sabine Lake and vicinity were not

highly contaminated. Chemical concentrations often were below respective ERL values. A

few chemical concentrations equalled or exceeded the ERL values and only one chemical

in one sample exceeded an ERM value. The spatial extent of chemical contamination was

restricted to a small percentage of the total survey area. Sediments from Taylor Basin and

a side-channel bayou of the Neches River had higher chemical concentrations than those

from other areas. Samples from the basin of Sabine Lake often were among the least

contaminated. Concentrations of organic compounds such as the PAHs decreased seaward and offshore into the Gulf of Mexico; however, the concentrations of trace metals such

as arsenic did not follow this pattern. Concentrations of arsenic in offshore samples were

equivalent to those from inland locations. Toxicity responses in the amphipod and urchin

fertilization tests were not significantly correlated with chemical concentrations; however,

results of urchin development and cytochrome P-450 RGS tests were highly correlated with

all chemical classes. Urchin development was very highly correlated with the concentrations of trace metals and porewater unionized ammonia and P-450 induction was highly

correlated with PAHs in the sediments.

Discussion and Conclusions

The survey of Sabine Lake and vicinity encompassed an area of approximately 246 km2

and included analyses of 66 sediment samples for chemical contamination and toxicity.

Samples from Sabine Lake, the Neches and Sabine rivers, the Intracoastal Waterway, and

the entrance channel (Sabine Pass) to the Gulf of Mexico were collected for analyses.

Toxicity was determined with a battery of acute and sublethal tests conducted on bulk

(solid-phase) sediments, pore waters, and organic solvent extracts. Chemical analyses

were conducted to determine the concentrations of many potentially toxic substances,

including ammonia, trace metals, polynuclear aromatic hydrocarbons, and chlorinated

organic compounds.

The results indicated that the sediments in this area were not significantly degraded. The

spatial extent of chemical contamination when compared to effects-based ERM values was

nil and the spatial extent when compared to ERL values was less than 1% for most substances. Chemical concentrations of organic substances were highest in Taylor Basin and

an arm of the Neches River near Beaumont; whereas concentrations of trace metals failed

to follow a clear spatial pattern. A very clear layer of oil was observed in the sediments

from an arm of Neches River below the upper 3 cm depth that was sampled, possibly the

result of a previous spill event. Overall, however, the sediments in the waterway channels

were more contaminated than those from the basin of Sabine Lake.
35

Table 9. A comparison of the spatial extent of toxicity (percentage of

total survey area) in amphipod survival tests among 25 estuarine

areas nationwide and "national estuarine averages" compiled from

field studies conducted through 1995, 1996, and 1997.

Total Amphipod survival

Estuarine area area (% of

Study Area (Km2

) (km2) total area)

 Newark Bay 13.0 10.8 85.0%

 San Diego Bay 40.2 26.3 65.8%

 California coastal lagoons 5.0 2.9 57.9%

 Tijuana River 0.3 0.18 56.2%

 Long Island Sound 71.9 36.3 50.5%

 Hudson-Raritan Estuary 350.0 133.3 38.1%

 San Pedro Bay 53.8 7.8 14.5%

 Biscayne Bay 484.2 62.3 12.9%

National average: 1995 2532.6 277.0 10.9%

National average: 1996 4158.1 286.4 6.9%

 Boston Harbor 56.1 5.7 10.0%

 Delaware Bay 2346.8 145.4 6.20%

National average: 1997 7278.8 431.8 5.93%

 Savannah River 13.1 0.16 1.2%

 St. Simons Sound 24.6 0.10 0.4%

 Tampa Bay 550.0 0.5 0.1%

 Galveston Bay 1351.1 0.0 0.0%

 northern Puget Sound 773.9 0.00 0.0%

 Pensacola Bay 273.0 0.04 0.0%

 Choctawhatchee Bay 254.5 0 0.0%

 Sabine Lake 245.9 0.00 0.0%

 Apalachicola Bay 187.6 0 0.0%

 St. Andrew Bay 127.2 0 0.0%

 Charleston Harbor 41.1 0 0.0%

 Winyah Bay 7.3 0 0.0%

 Mission Bay 6.1 0.0 0.0%

 Leadenwah Creek 1.7 0 0.0%

 San Diego River 0.5 0.0 0.0%

The estimated spatial extent of toxicity as determined in the amphipod survival test is

compared to comparable estimates for other estuarine regions studied by NOAA in Table 9.

The surficial area represented by toxic samples in Sabine Lake (i.e., none or 0.0% of the

area) was comparable with results from Galveston Bay, northern Puget Sound, Pensacola

Bay, Charleston Harbor and a number of other areas. Sabine Lake ranked well below the
36

Table 10. A comparison of the spatial extent of toxicity (percentage of 

total survey area) in urchin fertilization tests of 100% pore water among

22 estuarine survey areas and "national estuarine averages" compiled

from field studies conducted through 1995, 1996, and 1997.

Total Urchin fertilization @100%

Estuarine area area (% of

study area (Km2

) km 2 total area)

 San Pedro Bay 53.8 52.6 97.70%

 Tampa Bay 550 463.6 84.30%

 San Diego Bay 3 4 25.8 75.90%

 Mission Bay 6.1 4.0 65.90%

 Tijuana River 0.3 0.18 56.20%

 San Diego River 0.5 0.26 52.00%

 Biscayne Bay 484.2 229.5 47.40%

 Choctawhatchee Bay 254.47 113.14 44.40%

 California coastal lagoons 5 2.1 42.70%

National average: 1995 2082.6 886.3 42.60%

 Winyah Bay 7.3 3.1 42.20%

National average: 1996 3717.06 1439.9 38.67%

 Apalachicola Bay 187.58 63.6 33.90%

 Galveston Bay 1351.1 432.0 32.00%

 Charleston Harbor 41.1 12.5 30.40%

National average: 1997 6837.76 1728.0 25.25%

 Savannah River 13.12 2.42 18.40%

 Delaware Bay 2346.8 247.5 10.50%

 Boston Harbor 56.1 3.8 6.60%

 Sabine Lake 245.9 14.0 5.70%

 Pensacola Bay 273 14.4 5.30%

 northern Puget Sound 773.9 40.6 5.20%

 St. Simons Sound 24.6 0.65 2.60%

 St. Andrew Bay 127.2 2.28 1.80%

 Leadenwah Creek 1.69 0 0.00%

national estuarine averages calculated with data gathered through the 1995, 1996, and

1997 field seasons.

In the urchin fertilization tests performed with 100% pore water concentrations, Sabine

Lake also ranked among the least toxic areas (Table 10). The estimated spatial extent of

toxicity in Sabine Lake (6%) was comparable to that for Pensacola Bay (5%) and Boston

Harbor (7%) and less than the national estuarine averages for data compiled through the

1995, 1996, and 1997 field seasons.

In sharp contrast to the other tests, results of the urchin tests of embryological development

indicated toxic conditions throughout the area. The estimated spatial extent of toxicity in

Sabine Lake (99%) in the embryological development tests was much higher than in most

other areas studied by NOAA (Table 11). The degree of toxicity in Sabine Lake was only

slightly lower than in Boston Harbor (100% of the area) and higher than in Biscayne Bay

(84%) and in the national estuarine average (39%). The concentrations of the unionized

form of ammonia in the pore waters were sufficiently high in 35 of the 66 samples to cause
37

Table 11. A comparison of the spatial extent of toxicity 

(percentages of area) in urchin embryo development

tests performed with 100% pore water of sediments 

from 8 U.S. bays and estuaries.

Total Urchin development

area Toxic area (percent of

(Km2

) (km2) total area)

 Boston Harbor 56.1 56.8 100.0%

 Sabine Lake 246 245 99.4%

 Biscayne Bay 484.2 408 84.3%

 Apalachicola Bay 187.58 157.5 84.0%

 Choctawhatchee Bay 254.47 116.1 45.4%

National average 2734 1066 39.0%

 Galveston Bay 1351.1 314.8 23.3%

 St. Andrew Bay 127.2 7.2 5.6%

 Pensacola Bay 273 5.4 2.0%

or contribute significantly to decreased embryological development; thus, possibly leading -

unlike the results of the other tests - to the observation of widespread toxicity in this test.

Results of Microtox tests are compared among study areas in Table 12. Using 80% of

controls as the critical value in these calculations, the spatial extent of toxicity in Sabine

Lake (79%) was greater than comparable results from most other areas and the national

estuarine averages. However, comparisons to the Redfish Bay negative controls may

exaggerate the degree of toxicity in Sabine Lake. Using a critical value of <0.51 mg/ml (as

done in the northern Puget Sound and Delaware Bay studies) provides an estimated spatial

extent of toxicity of 1.4% in Sabine Lake. This estimate puts Sabine Lake toward the bottom of the list and is comparable with the estimates for the other two regions.

Comparable data from the cytochrome P450 assays are available from five other areas

(Table 13). Estimates of the spatial extent of significant induction (i.e., critical value of 11.1

ug/g) and high induction (i.e., >37.1 ug/g) in Sabine Lake were somewhat lower than in

other areas. Results were comparable with those from northern Puget Sound.

Based upon the compilation of results from chemical analyses and toxicity tests of surficial

samples from 66 locations, sediments in Sabine Lake and vicinity did not appear to be

severely degraded. Chemical concentrations rarely exceeded effects-based numerical

guidelines, suggesting that toxicant-induced effects would not be expected in most areas.

None of the samples was highly toxic in acute amphipod survival tests and a minority of

samples was highly toxic in sublethal urchin fertilization tests. Although toxic responses

occurred frequently in urchin embryo development tests performed with 100% pore water,

toxicity diminished frequently in tests done with diluted pore waters. Microbial bioluminescent activity was not reduced to a great degree and cytochrome P-450 activity was not

highly induced in tests done with organic solvent extracts. Urchin embryological development was highly correlated with concentrations of ammonia and many trace metals. Cytochrome P-450 induction was highly correlated with concentrations of a number of classes

of organic compounds.
38

Table 12. Spatial extent of toxicity (km2 and percentages of total area) in microbial

bioluminescence tests performed with solvent extracts of sediments from 17 U. S.

bays and estuaries.

Total Microtox bioluminescence

Estuarine area Area (% of

study area (Km2

) km 2 total area)

 Choctawhatchee Bay 254.47 254.47 100.00%

 St. Andrew Bay 127.2 127 100.00%

 Apalachicola Bay 187.58 186.84 99.60%

 Pensacola Bay 273 262.8 96.40%

 Galveston Bay 1351.1 1143.7 84.60%

 Sabine Lake 245.9 194.2 79.00%

 Winyah Bay 7.3 5.13 69.99%

 Long Island Sound 71.86 48.8 67.90%

National average: 1996 4039.22 2670.69 66.12%

National average: 1995 2416.2 1482.3 61.30%

 Savannah River 13.12 7.49 57.10%

 Biscayne Bay 484.2 248.4 51.30%

 St. Simons Sound 24.6 11.42 46.40%

 Boston Harbor 56.1 25.8 44.90%

 Charleston Harbor 41.1 17.6 42.90%

National average: 1997 7159.92 2802.39 39.10%

 Hudson-Raritan Estuary 350 136.1 38.90%

 Leadenwah Creek 1.69 0.34 20.10%

 Delaware Bay (<0.51 mg/ml) 2346.8 114 4.90%

 northern Puget Sound (<0.51 mg/ml) 773.9 17.7 2.20%

 Tampa Bay 550 0.6 0.10%

Table 13. Spatial extent of toxicity (km2 and percentages of total area) in P-450

RGS bioassays performed with solvent extracts of sediments from 6 areas.

Total P-450 RGS (>11.1 ug/g) P-450 RGS (>37.1 ug/g)

area (% of (% of

(Km2

) km 2 total area) km 2 total area)

Biscayne Bay, 1996 271.4 8.8 3.3 0 0

northern Puget Sound 806.2 20.1 2.5 0.2 0.03

Delaware Bay 2346.8 145.2 6.2 80.5 3.4

Galveston Bay 1351.5 56.7 4.2 0 0

Sabine Lake 245.9 6.7 2.7 1.7 0.7

Southern Cal. Estuaries 5 2.30 46.8 0 0
39

Acknowledgments

Field logistics were conducted aboard the NOAA vessel Ferrell and its launch. The officers

and crew of the Ferrell were excellent, working hard and cheerfully in very hot and trying

conditions. The Coast Guard provided excellent logistic support and cooperation by allowing us to dock the vessel at the Sabine Pass station. Miss Heather Boswell provided assistance by determining sampling station coordinates and preparing base maps of the study

area. EVS Environment Consultants (Corinne Severn, Carolyn Hong) assisted in preparation of data spreadsheets.

References

ASTM. 1992. Standard Guide for Conducting 10-day Static Toxicity Test with Marine and

Estuarine Amphipods. Designation E 1367-92. Annual Book of Standards. 11.04. American Society for Testing and Materials. Philadelphia, PA.

ASTM. 1996. Standard guide for measuring the presence of planar organic compounds

which induce CYP1A, using reporter gene test systems. E 1853-96. American Society for

Testing Materials. Philadelphia, PA. 1392-1397.

ASTM. 1997. E 1853 standard guide for measuring the presence of planar organic

compunds which induce CYP1A, reporter gene test systems. Vol. 1.05. American Society

for Testing and Materials. Philadelphia, PA. In press.

Anderson, J. W., S. S. Rossi, R. H. Tukey, T. Vu, and L. C. Quattrochi. 1995. A biomarker,

450 RGS, for assessing the potential toxicity of organic compounds in environmental

samples. Envir. Toxicol. & Chem. 14(7): 1159-1169.

Anderson, J. W., K. Bothner, T. Vu, and R. H. Tukey. 1996. Using a biomarker (P450 RGS)

test method on environmental samples. pp. 277-286. Chapter 15 in: Techniques in Aquatic

Toxicology. Editor - G. K. Ostrander. Lewis Publishers, Boca Raton, FL.

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25. In: Standard Methods for the Examination of Water and Wastewater. 19th edition

supplement. American Public Health Association, Washington, D.C.

Carr, R.S. and D.C. Chapman. 1992 Comparison of solid-phase and pore-water approaches for assessing the quality of marine and estuarine sediments. Chem. Ecol. 7:19-

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Johnson, B. T. and E. R. Long. 1998. Rapid Toxicity Assessment of Sediments from Estuarine Ecosystems: A New Tandem In Vitro Testing Approach. Environmental Toxicology and

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Hamilton, M. A. , R. C. Russo, and R. V. Thurston. 1977. Trimmed Spearman-Karber

method for estimating median lethal concentrations in toxicity bioassays. Environ. Sci.

Technol. 11: 714-719.
40

Harrel, R.C. and M. A. McConnell. 1995. The estuarine clam Rangia cuneata as a

biomonitor of dioxins and furans in the Neches River, Taylor Bayou, and Fence Lake,

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Heimbuch, D.G. H.T. Wilson, D.S. Robson, and J.K. Summers. 1995a. Design-based estimation of the proportion of area degraded and asociated variance estimators for EMAPEstuaries sampling in the Louisianian andVirginian Provinces: 1990-1994. Draft report

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Heimbuch, D.G. H.T. Wilson, D.S. Robson, and J.K. Summers. 1995b. Power analysis for

tests for long term trends in environmental conditions of estuaries. Draft report prepared for

U.S. Environmental Protection Agency by Coastal Environmental Services, Linthicum,

Maryland.

Kohn, N. P., J.Q. Word, D. K. Niyogi. 1994. Acute Toxicity of ammonia to four species of

marine amphipod. Mar. Env. Res. 38 (1994) 1-15.

Lauenstein, G. G. and A. Y. Cantillo, editors. 1993. Sampling and analytical methods of

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Atmopheric Administration. Silver Spring, MD.

Lauenstein, G. G. and A. Y. Cantillo, editors. 1998. Sampling and analytical methods of

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Long, E. R., D. D. MacDonald, S. L. Smith, and F. D. Calder. 1995. Incidence of adverse

biological effects within ranges of chemical concentrations in marine and estuarine sediments. Environmental Management 19(1): 81-97.

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the spatial extent of sediment toxicity in major U. S. estuaries. Environmental Science &

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Ravichandran, M., M. Baskaran, P. H. Santschi, and T. S. Bianchi. 1995. History of trace

metal pollution in Sabine-Neches estuary, Beaumont, Texas. Environ. Sci. Technol. 29:

1495-1503.

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vol. 2. Cary, NC: SAS Institute, Inc. 846 pg.
41

Schimmel, S. C., B. D. Melzian, D. E. Campbell, C. J. Strobel, S. J. Benyi, J. S. Rosen, and

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Chemistry 16(6): 1322-1329.
42

Appendix A. Field notes on sampling stations in Sabine Lake and vicinity.

Location Stratum No.Station No. Sample Id. Latitude °N Longitude °W Depth (ft.) Depth(m) Date

Neches River

A 1 1 30°05.5' 94°05.27' 34.3 10.46 8/8/95

2 2 30 ̊04.4833' 94 ̊03.55' 39.5 12.04 8/17/95

3 3 30 ̊02.75' 94 ̊01.7' 47.3 14.42 8/17/95

B 1 4 30 ̊01.18' 94 ̊01.3' 47.2 14.39 8/17/95

2 5 30 ̊00.624' 93 ̊59.282' 47.7 14.54 8/17/95

3 6 29 ̊59.910' 93 ̊57.040' 47.7 14.54 8/17/95

C 1 7 29 ̊59.512' 93 ̊56.084' 47.8 14.57 8/16/95

2 8 29 ̊59.097' 93 ̊53.806' 41.5 12.65 8/16/95

3 9 29 ̊58.879' 93 ̊52.421' 38.5 11.74 8/16/95

D 1 10 30 ̊04.72' 94 ̊03.167' 49.1 14.97 8/18/95

2 11 30 ̊03.917 94 ̊02.00' 11.3 3.45 8/18/95

3 12 30 ̊02.617' 94 ̊02.15' 28.5 8.69 8/18/95

Sabine River

E 1 13 30 ̊02.833 93 ̊43.75' 16.1 4.91 8/15/95

2 14 30 ̊01.2' 93 ̊44.833' 22 6.71 8/15/95

3 15 29 ̊59.9' 93 ̊46.42' 31.3 9.54 8/15/95

F 1 16 29 ̊59.4' 93 ̊47.37' 32.8 10.00 8/15/95

2 17 29 ̊59.067' 93 ̊48.5' 30.5 9.30 8/15/95

3 18 29 ̊58.38' 93 ̊50.7' 22.8 6.95 8/15/95
43

Appendix A. Field notes on sampling stations in Sabine Lake and vicinity continued.

Location Stratum No.Station No. Sample Id. Latitude °N Longitude °W Depth (ft.) Depth(m) Date

Sabine-Neches Canal

G 1 19 29 ̊57.937 93 ̊51.363 37.1 11.31 8/16/95

2 20 29 ̊56.801' 93 ̊51.820 42.3 12.90 8/16/95

3 21 29 ̊55.35 93 ̊52.917 32.5 9.91 8/14/95

H 1 22 29 ̊54.6' 93 ̊53.53 35 10.67 8/14/95

2 23 29 ̊53.98' 93 ̊54.017' 34.8 10.61 8/14/95

3 24 29 ̊53.017' 93 ̊55.083' 38 11.59 8/14/95

I 1 25 29 ̊52.383' 93 ̊55.73' 41.5 12.65 8/14/95

2 26 29 ̊51.083' 93 ̊56.82' 32 9.76 8/14/95

3 27 29 ̊49.53' 93 ̊57.467' 10.5 meters 10.50 8/8/95

J 1 28 29 ̊50.58 * 93 ̊58.13 * 44 13.41 8/9/95

2 29 29 ̊50.22 * 93 ̊58.067'* 44 13.41 8/9/95

3 30 29 ̊50.05' 93 ̊57.45' 41 12.50 8/9/95

K 1 31 29 ̊48.383' 93 ̊57..267' 12 meters 12.00 8/8/95

2 32 29 ̊47.48 93 ̊56.933' 12 meters 12.00 8/8/95

3 33 29 ̊45.933' 93 ̊55.17' 12 meters 12.00 8/8/95

L 1 34 29 ̊44.872 93 ̊.52.936 51.8 15.79 8/10/95

2 35 29 ̊43.611 93 ̊51.941 47.8 14.57 8/10/95

3 36 29 ̊42.148 93 ̊50.881 49.3 15.03 8/10/95
44

Appendix A. Field notes on sampling stations in Sabine Lake and vicinity continued.

Location Stratum No.Station No. Sample Id. Latitude °N Longitude °W Depth (ft.) Depth(m) Date

Sabine Lake

M-1 1 37 29 ̊58.55' 93 ̊48.23' 5.1 1.55 8/19/95

2 38 29 ̊57.2' 93 ̊50.183' 5.2 1.59 8/19/95

3 39 29 ̊58.85' 93 ̊47.9 3.9 1.19 8/19/95

M-2 1 40 29 ̊56.82 93 ̊48.967 7.5 2.29 8/19/95

2 41 29 ̊56.7' 93 ̊47.467' 7.6 2.32 8/19/95

3 42 29 ̊55.433 93 ̊51.7 5.9 1.80 8/19/95

M-3 1 43 29 ̊52.533' 93 ̊48.82' 8.4 2.56 8/21/95

2 44 29 ̊54.67' 93 ̊50.733' 7.5 2.29 8/21/95

3 45 29 ̊54.883' 93 ̊51.87' 6 1.83 8/21/95

M-4 1 46 29 ̊52.402 93 ̊50.921 7.7 2.35 8/12/95

2 47 29 ̊52.446 93 ̊47.368 8.2 2.50 8/12/95

3 48 29 ̊51.263 93 ̊49.320 8.4 2.56 8/12/95

M-5 1 49 29 ̊52.405 93 ̊51.710 7.3 2.23 8/13/95

2 50 29 ̊50.32' 93 ̊54.92' 6.1 1.86 8/13/95

3 51 29 ̊50.7' 93 ̊54.33' 6.4 1.95 8/13/95

M-6 1 52 29 ̊48.85' 93 ̊53..333' 6.5 1.98 8/13/95

2 53 29 ̊50.67' 93 ̊54.4' 6.1 1.86 8/13/95

3 54 29 ̊49.65' 93 ̊52.96' 6.9 2.10 8/13/95

M-7 (a) 1 55 29 ̊48.036 93 ̊55.038 8.5 2.59 8/12/95

2 56 29 ̊46.887 93 ̊55.065 10.3 3.14 8/12/95

3 57 29 ̊47.699 93 ̊55.808 6.7 2.04 8/12/95

Sabine Pass

O 1 58 29 ̊41.098 93 ̊50.144 55.5 16.92 8/10/95

2 59 29 ̊39.425 93 ̊49.738 45 13.72 8/10/95

3 60 29 ̊37.476 93 ̊48.833 29.1 8.87 8/11/95

P 1 61 29 ̊38.62' 93 ̊49.87' 17.6 5.37 8/11/95

2 62 29 ̊37.005' 93 ̊50.22' 20.9 6.37 8/11/95

3 63 29 ̊37.6' 93 ̊50.233' 20.1 6.13 8/11/95

R 1 64 29 ̊38.708 93 ̊48.898 30 9.15 8/11/95

2 65 29 ̊37.626 93 ̊48.507 27.4 8.35 8/11/95

3 66 29 ̊37.803 93 ̊48.359 27.4 8.35 8/11/95

* No coordinates obtained at the station, target coordinates are listed. 
45