Notes from Rao 2/13/2007 2:07:37 PM

Environmental Assessment of Sediments and Water in Bayou Grande, Pensacola, FL.

Dr. Carl J. Mohrherr Center for Environmental Diagnostics and Bioremediation University of West Florida

Dr. Johan Liebens Department of Environmental Studies University of West Florida

Dr. K. Ranga Rao Center for Environmental Diagnostics and Bioremediation University of West Florida

June 26, 2008 FOREWORD

This study is a component of the "Assessment of Environmental Pollution and Community Health in Northwest Florida" supported by a USEPA Cooperative Agreement award X-9745502 to The University of West Florida (Project Director: Dr. K. Ranga Rao). The contents of this report are solely the responsibility of the authors and do not necessarily represent the official views of the USEPA. The study was undertaken because of the increasing concern for environmental pollution and potential impacts on human health in Northwest Florida. It was designed to assess environmental impacts of toxic pollutants in Bayou Grande. Kristal Flanders managed the spatial databases for the project and drafted the maps. Her assistance has been invaluable. Chris Carlton-Franco, Brandon Jarvis, Guy Allard, and Danielle Peterson helped with the fieldwork and some laboratory procedures.

Creative Commons License This work is licensed under a Creative Commons Attribution- NonCommercial- NoDerivatives 4.0 International License. i TABLE OF CONTENTS

I INTRODUCTION...... 1

II STUDY AREA...... 2

II.1 Physiography...... 2

II.2 Climate...... 3

II.3 Urban Development...... 4

II.4 Pensacola Bay and Dredging Activity...... 5

II.5 Historical Outline of Bayou Grande ...... 6

III POLLUTION IN BAYOU GRANDE: LITERATURE REVIEW...... 8

III.1 General Pollution History...... 8

III.2 Fecal Coliform Pollution in Bayou Grande ...... 13

III.3 Total Petroleum Hydrocarbons ...... 14

III.4 Polycyclic Aromatic Hydrocarbons (PAHs) ...... 15

III.5 Organochlorinated Compounds...... 20

III.6 Trace Metals ...... 21

IV ENVIRONMENTAL BACKGROUND OF SEDIMENT POLLUTANTS ...... 26

IV.1 Polychlorinated Biphenyls (PCBs)...... 26

IV.2 Organochlorinated Compounds...... 28 IV.2.1 General Notions ...... 28 IV.2.2 Organochlorinated Pesticides in Sediments and Biota...... 28 IV.2.2.1 Aldrin and Dieldrin...... 28 IV.2.2.2 Endrin ...... 29 IV.2.2.3 Lindane ...... 30 IV.2.2.4 Chlordane...... 31 IV.2.2.5 DDT ...... 32 IV.2.2.6 Mirex and Chlordecone ...... 33 IV.2.2.7 Endosulfan ...... 36 IV.2.3 Dioxins/Furans...... 37 IV.2.4 Pentachlorophenol (PCP) ...... 39

IV.3 Metals...... 40 IV.3.1 Mercury (Hg) ...... 40 IV.3.2 Lead (Pb) ...... 42 IV.3.3 Cadmium (Cd) ...... 44 IV.3.4 Arsenic (As)...... 46 IV.3.5 Chromium (Cr)...... 47 IV.3.6 Copper (Cu) ...... 49 IV.3.7 Nickel (Ni)...... 50 IV.3.8 Zinc (Zn)...... 51

V OBJECTIVES ...... 53

VI METHODS...... 54

VII RESULTS AND DISCUSSION ...... 60

VII.1 Semivolatile Organic Compounds...... 60 VII.1.1 Total Petroleum Hydrocarbons...... 60 VII.1.2 Polycyclic Aromatic Hydrocarbons (PAHs)...... 63 VII.1.2.1 PAHs in surface sediments...... 63 VII.1.2.2 PAHs in the water column...... 74 VII.1.2.3 PAHs in vibracores...... 74 VII.1.2.4 Origins of PAHs in Bayou Grande...... 80 VII.1.2.5 Pentachlorophenol (PCP) ...... 87

VII.2 Dioxins/furans and PCBs ...... 88 VII.2.1 Dioxin/furan and PCB TEQ...... 88 VII.2.2 Dioxin/furan and PCB TEQ in Bayou Grande Sediments...... 88 VII.2.3 Total Dioxin/Furan Mass Concentrations in Surface Sediments ...... 92 VII.2.4 Total Dioxin/Furan Mass Concentrations in Vibracores...... 93 VII.2.5 Origin of Dioxins/Furans in Bayou Grande ...... 94 VII.2.6 Total PCB Concentrations in Surface Sediments ...... 98 VII.2.7 Total PCB Concentrations in Vibracores...... 101 VII.2.8 Origin of PCBs in Bayou Grande...... 101 VII.2.9 Dioxin-like PCB Concentrations in Surface Sediments ...... 109 VII.2.10 Degradation of PAHs, PCBs, and Dioxins/ Furans in Sediments...... 113 VII.2.11 Dioxins/Furans in Bayou Grande Sediments and Seafood Tissues...... 120 VII.2.12 PCBs in Bayou Grande Sediments and Seafood Tissues...... 123

VII.3 Pesticides...... 126

VII.4 Trace Metals...... 129 VII.4.1 Trace Metal Concentrations in Surface Sediments ...... 129 VII.4.2 Origin of Trace Metals in Surface Sediments ...... 140 VII.4.3 Trace Metal Concentrations in Vibracores ...... 144 VII.4.4 Trace Metal Concentrations in Water ...... 147

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VIII CONCLUSIONS ...... 150

IX REFERENCES ...... 153

LIST OF FIGURES

Figure 1. Location of Bayou Grande...... 3 Figure 2. Bayou Grande and nearby drainage basins ...... 4 Figure 3. Dredged shipping channels in Pensacola Bay...... 5 Figure 4. Map of Pensacola Bay, 1860s (Exploring Florida, 2008a)...... 7 Figure 5. Sketch of Pensacola Navy Yard and part of Fort Pickens, 1860s (Exploring Florida, 2008b)...... 8 Figure 6. Location of Assessment Zones at NAS...... 11 Figure 7. Geomeans of Enterococcus at zero rainfall for Bayou Grande...... 14 Figure 8. Location of sampling sites from DeBusk et al. (2002) GIS database...... 18 Figure 9. General PCB structure...... 26 Figure 10. Structure of Aldrin...... 29 Figure 11. Structure of Dieldrin...... 29 Figure 12. Structure of Endrin...... 30 Figure 13. Structure of Lindane (gamma-hexachlorocyclohexane)...... 31 Figure 14. Structure of Chlordane...... 31 Figure 15. Structure of DDT (di-chlorodiphenyltrichloroethane)...... 32 Figure 16a. Structure of Mirex...... 34 Figure 16b. Structure of Chlordecone...... 34 Figure 17. Structure of Endosulfan...... 36 Figure 18. Chemical Structure of 2,3,7,8-dioxin (TCDD) and representative dioxin-like compounds...... 38 Figure 19. Cycling of mercury between the atmosphere, water, sediments, and organisms (Stein et al., 1996)...... 42 Figure 20. Location of water grab samples...... 56 Figure 21. Location of vibracore sites in Bayou Grande...... 56 Figure 22. Total petroleum hydrocarbons in sediments...... 62 Figure 23. Total PAH concentration in sediments...... 71 Figure 24. Sum of 13 LMW and HMW PAHs in sediments...... 71 Figure 25. Total naphthalenes in surface sediments...... 81 Figure 26. Combined TEQ of dioxins/furans and PCBs...... 90 Figure 27. Spatial distribution of TEQ for dioxins/furans...... 91 Figure 28. Spatial distribution of TEQ for dioxin-like PCBs...... 91 Figure 29. Total dioxin/furan mass concentrations in surface sediments...... 93 Figure 30. Factor loading plot for first two principal components for dioxins/furans - all profiles...... 96 Figure 31. Factor loading plot for first two principal components for dioxins/furans - selected profiles...... 97 Figure 32. Average dioxin/furan homologue profile for GV samples...... 98 Figure 33. Total PCB concentrations in surface sediments...... 100 Figure 34. Factor loading plot for first two principal components for PCBs ...... 104 Figure 35. Cluster analysis of PCB congener data for surface sediments...... 105 Figure 36a. PCB homologue profiles for GF series samples and Aroclors...... 106 Figure 36b. PCB homologue profiles for surface sediments from GV series and Aroclors...... 107

Figure 36c. PCB homologue profiles for subsurface sediments from GV series and Aroclors. 108 Figure 37. PCB homologues in vibracore GV-1...... 115 Figure 38. PCB homologues in vibracore GV-4...... 116 Figure 39. PCB homologues in vibracore GV-11...... 116 Figure 40. Ortho, meta, and para positions on aromatic molecules...... 117 Figure 41. Dechlorination of dioxin/furan congeners...... 120 Figure 42. Dioxins/furans in Bayou Grande seafood tissues...... 121 Figure 43. Dioxins/furans in Bayou Grande crab hepatopancreas and sediment...... 122 Figure 44. Dioxins/furans in Bayou Grande oysters and sediment...... 123 Figure 45. Dioxin-like PCBs in seafood tissues and sediment...... 124 Figure 46. Dioxin-like PCBs in crab hepatopancreas tissue and sediment...... 125 Figure 47. Dioxin-like PCBs in Bayou Grande oyster tissue and sediment...... 125 Figure 48. Arsenic in Bayou Grande sediments...... 130 Figure 49. Cadmium in Bayou Grande sediments...... 131 Figure 50. Chromium in Bayou Grande sediments...... 135 Figure 51. Copper in Bayou Grande sediments...... 136 Figure 52. Lead in Bayou Grande sediments...... 137 Figure 53. Mercury in Bayou Grande sediments...... 138 Figure 54. Nickel in Bayou Grande sediments...... 139 Figure 55. Zinc in Bayou Grande sediments...... 140 Figure 56. Factor loading plot for first three components for trace metals in surface sediments...... 141 Figure 57. Spatial distribution of average standardized pollution index (ASPI) for trace metals in surface sediments...... 144

LIST OF TABLES

Table 1. Sources of data in environmental GIS database for Pensacola Bay System (DeBusk et al., 2002)...... 12 Table 2. Chemical and physical requirements for JP-4, JP-5, and JP-8...... 15 Table 3. PAH SQAGs1 (μg/kg) and IARC listing ...... 16 Table 4. Existing ‘non-NAS’ Bayou Grande PAH data [ug/kg] (DeBusk et al., 2002) ...... 17 Table 5. PAHs in Bayou Grande sediments by zone of the NAS Pensacola shore (EnSafe, 2003; 2004) ...... 19 Table 6. Existing pesticide data for Bayou Grande [ug/kg] (DeBusk et al., 2002)...... 20 Table 7. Pesticides and PCB Aroclors in Bayou Grande sediments by zone of the NAS Pensacola shore (EnSafe, 2003; 2004)...... 22 Table 8. Existing total PCB data for Bayou Grande [ug/kg] (DeBusk et al., 2002)...... 23 Table 9. Water quality standards [ug/l] for Class III marine water bodies1...... 23 Table 10. Existing data for metals in sediments [mg/kg] in Bayou Grande (DeBusk et al., 2002)...... 24 Table 11. Metals in Bayou Grande sediments by zone of the NAS Pensacola shore...... 25 Table 12. Major PCB congener constituents of five Aroclors [%]...... 27 Table 13. TEF values for dioxins/furans and dioxin-like PCBs ...... 59 Table 14. Total petroleum hydrocarbons [mg/kg] in surface sediments...... 61 Table 15. Total petroleum hydrocarbon [mg/l] in water...... 62 Table 16. Typical PAH minimal detection limit (MDL) and reporting limit (RL) and SQAGs (from sample GBc-30)...... 64 Table 17. PAH concentrations [ug/kg] in sediments, GBc series...... 67 Table 18. PAH concentrations [ug/kg] in sediments, GF series...... 72 Table 19. PAHs and pentachlorophenol in water samples [ug/l]...... 75 Table 20. Total PAHs in vibracores by depth level [ug/kg]...... 76 Table 21. Specific PAHs in Bayou Grande vibracores [ug/kg]...... 77 Table 22. Naphthalene concentrations [ug/kg] in sediments, GF series...... 82 Table 23. Naphthalene concentrations [ug/kg] in sediments, GBc series...... 83 Table 24. PAH origin indicator ratios, Yunker et al. (2002)...... 83 Table 25. PAH ratios in sediments, GF series...... 85 Table 26. PAH ratios in sediments, GBc series...... 86 Table 27. TEQ for dioxins/furans, dioxin-like PCBs, and combined total TEQ [ng/kg]...... 89 Table 28. Dioxin/furan mass and TEQ concentrations [ng/kg] in surface sediments...... 92 Table 29. Dioxin/furan mass and TEQ concentrations [ng/kg} in sediment from vibracores...... 94 Table 30. Average dioxin/furan congener composition for Bayou Grande surface sediments. ... 95 Table 31. Total PCB concentrations [ug/kg] in surface sediments...... 99 Table 32. Total PCB concentrations [ug/kg] for vibracores...... 102 Table 33. Average PCB congener profile for surface sediment in Bayou Grande...... 103 Table 34. Dioxin-like PCB TEQ [ng/kg] and total PCB mass concentration [ug/ kg] in surface sediments...... 109 Table 35. Dioxin-like PCB congener mass concentration [ng/kg] for surface sediments...... 111 Table 36. Dioxin-like PCB congener TEQs [ng/kg] for surface sediments ...... 112 Table 37. Average dioxin-like PCB mass concentration and TEQ...... 113

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Table 38. Dioxin-like PCB congener distributions [%] for seven Aroclor mixtures (modified from Frame et al., 1996)...... 113 Table 39. Selected vibracore PCB, PCB TEQ, and PAH concentrations...... 114 Table 40. TEQ-to-mass ratio for dioxins/furans in vibracore samples...... 118 Table 41. TEFs for dioxin/furan congeners...... 119 Table 42. Dioxins/Furans in seafood tissues [ng/kg wet wt] and sediments [ng/kg dry weight]...... 121 Table 43. Dioxin-Like PCBs in crab and oyster tissues [ng/kg wet wt] and sediment [ng/kg dry weight]...... 123 Table 44. Organochlorinated pesticide concentrations in surface sediments [ug/kg]...... 127 Table 45. Organochlorinated pesticide concentrations in water samples [ug/l]...... 129 Table 46. Trace metals in surface sediments [mg/kg]...... 132 Table 47. Average standardized pollution index for anthropogenic trace metals...... 142 Table 48. Trace metals in vibracores [mg/kg]...... 145 Table 49. Aqueous metal concentrations in Bayou Grande [ug/l]...... 148 Table 50. Water parameters for Bayou Grande, February 2007...... 149 Table 51. Water parameters determined in laboratory...... 149

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EXECUTIVE SUMMARY

The PERCH (Partnership for Environmental Research and Community Health) Project component on Bayou Grande was designed to address community concerns relating to environmental health issues for this locally important water body. Bayou Grande is the largest of the three urban bayous in the Pensacola metropolitan area. Pollutants affecting the water and sediment quality of the southern half of the Bayou have been studied in reference to possible releases from the Naval Air Station (NAS) on its southern shore, but not in reference to the entire bayou system. High levels of substances of concern (SOCs) including trace metals, polycyclic aromatic hydrocarbons (PAHs), dioxins/furans and polychlorinated biphenyls (PCBs) have been found in the sediments. The State of Florida classifies Bayou Grande as a 3M body of water that is suitable for recreational uses and for the propagation of fish and wildlife.

Human activities have likely affected the bayous in the area, including Bayou Grande, from the time European settlers entered the area. Initial impacts would have been caused by land clearing for agricultural and logging activities. US Naval operations in the Bayou Grande area began in the early 19th century. State of Florida scientists came to Pensacola in the 1950s to study pollution in the local bayous but focused less on Bayou Grande than on Bayous Chico and Texar. In the second half of the 20th century, however, evidence for pollution was observed in Bayou Grande due to increasing activities at NAS and increasing urbanization on the north shore of the Bayou in the Warrington area, as well as import from Pensacola Bay with which Bayou Grande connects. Among other pollutants were elevated fecal coliforms, as observed by other monitoring studies.

The present study examined preexisting environmental databases and publications, some of which were incorporated into a geographic information system (GIS). The existing information was utilized in prioritizing research efforts based on perceived gaps in the information. The main gap is a scarcity of systematic information about the environmental quality of the northern half of the Bayou. Parts of the Bayou along NAS have been studied to some extent, but information about sediment quality in northern sections of the Bayou and the Bayou's northern embayment is especially scarce. The presence of aviation and maritime activities at NAS suggests that petroleum contamination may be present in the Bayou, but this has not been fully addressed by previous studies. PERCH project found high levels of dioxins/furans in nearby Bayou Chico, but dioxin/furan analysis do not appear to have been carried out in Bayou Grande, in spite of their potential health impacts. The finding of elevated dioxin/furan and PCB levels in seafood from Bayou Grande warrants a more detailed analysis of these pollutants in the Bayou’s water and sediments. In addressing these knowledge gaps, the present study is the first to systematically study a large suite of sediment pollutants for the whole Bayou and to interpret results in relation to potential human health effects.

The fieldwork for the project took place from Fall 2006 to mid-summer 2007. The collected samples include: 78 sediment grab samples, collected with a ponar grab sampler; 8 water grabs, collected with a Van Dorn sampler; and 10 vibracores obtained with an in-house built vibracore system operated from a pontoon boat. The sediment grab samples consisted of two series, one series of 55 samples along the shoreline and the other of 23 samples in the channel of the Bayou

and its main embayments, including one sample in Pensacola Bay just outside Bayou Grande that served as a reference station. Chemical analyses were performed by a contracted commercial lab according to USEPA and Florida Department of Environmental Protection standards. Physical analyses of the sediments were conducted at UWF.

For total petroleum hydrocarbons (TPH) there were four detections in 23 sediment samples and three detections in eight water samples. All detected concentrations were very low and were below the reporting limit for the method. This indicates that although the potential for TPH contamination exists in Bayou Grande, the TPH levels remain low in the Bayou. Previous PERCH studies at nearby bayous (Bayous Chico and Texar) found more detections and higher concentrations of TPH.

PAHs are among the most toxic components of petroleum products, and are also associated with carcinogenic effects. All 18 PAHs detected by the USEPA 8270C method were found in Bayou Grande. The highest total PAH concentrations were observed in sediments in the Yacht Basin and other embayments of the eastern portion of the main body of the Bayou. Florida sediment quality guidelines (TEL - concentrations above which adverse effects on biota are possible, and PEL - concentrations above which adverse effects on biota are probable) were exceeded in these embayments by several of the PAH species. Concentrations of most PAH species decreased abruptly with depth in the vibracores. In water, PAH concentrations were generally low. The naphthalene content of the PAHs in the sediments suggests that the PAHs may be of petroleum origin but they also exhibit characteristics consistent with other origins such as combustion and coal tar. Naphthalenes have been detected in NAS groundwater indicating that transport by contaminated groundwater to the bayou is possible.

The toxicity of dioxins/furans and dioxin-like PCBs is expressed as Toxic Equivalents (TEQ). To determine the combined TEQ, all toxic dioxins/furans and dioxin-like PCBs have been assigned a Toxic Equivalency Factor (TEF) as defined by Van den Berg et al. (2006). In Bayou Grande, 17 of the 23 samples had a combined TEQ that exceeded the NOAA sediment quality guideline (AET). Seven of these samples had a TEQ almost three times the NOAA AET. The highest concentrations were found in two embayments along the southern shoreline and in an embayment of the northern shore, but exceedances occurred throughout the Bayou.

Principal Component Analysis of the congener profiles of the dioxins/furans indicates multiple origins. Samples from the deeper parts of the Bayou and its embayments have a profile that is consistent with a PCP origin. PCP was not detected in the sediments and we know of no historical wood treating activities near Bayou Grande that used PCP but it is possible that treated wood with PCP has been employed along the Bayou Grande shoreline, or that PCP was used in antifouling paints for boats. Other samples have dioxin/furan congener profiles that indicate an origin from forest fires and burning of oil in industrial boilers. Samples from deeper sediment levels have a clearly different congener profile that is consistent with an origin from effluent of wastewater treatment facilities. However, we have no indication from existing literature that such effluent was released to the Bayou even though industrial wastewater was. An alternative explanation is that dechlorination of the dioxins/furans in the deeper sediments by chance created a profile similar to that of wastewater. Concentrations of dioxins/furans in the deeper sediments are only slightly below those in surface sediments.

Homologue patterns of the PCBs in many of the surface sediment samples are similar to those of Aroclor 1260. Samples from the northern shoreline and northern embayments have homologue patterns consistent with a mixture of Aroclor 1260 and Aroclor 1254. Aroclor 1260 has been used at NAS, and 1254 is one of the most widespread Aroclors. These observations, together with the spatial distribution of the PCBs, suggest that much of Bayou Grande may be influenced by PCBs originating from NAS and that along the northern shore, near residential areas such as Warrington, PCBs from other sources have been mixed in.

Dioxins/furans and PCBs in sediments can enter into the food chain and upon entering seafood are consumed by humans with probable impacts upon human health. In Bayou Grande, profiles of dioxin-like PCB congeners for crab hepatopancreas, crab muscle, and oysters are very similar to profiles for sediments. This indicates that dioxin-like PCBs are bioaccumulating in seafood in Bayou Grande and that the accumulation is proportional to the relative congener composition of the PCBs in the sediments. Comparison between profiles of dioxin/furan congeners shows similarities between profiles for sediments and oysters but not for sediments and crabs. This suggests that crabs are either selective in the specific congeners that are incorporated into their tissues or that dechlorination is occurring once the congeners are incorporated into the crab. It appears that incorporation of dioxins/furans in oysters, but not in crabs, is related to the relative proportions of the dioxin/furan congeners in the sediments. Similar observations were reported in a previous PERCH study of nearby Bayou Chico.

Only five detections of organochlorinated pesticides occurred in the 23 sediment grab samples from Bayou Grande. These five detections occurred in only two samples in the mid section of the Bayou. One of the detections was for DDT and exceeded the TEL. There were more detections of organochlorinated pesticides in the 1990s. Most of these pesticides have not been applied for many years and our results suggest that the concentrations are declining in sediments of Bayou Grande. Sediment transport into and out of the Bayou can in principle reduce the concentrations of the pesticides as can abiotic degradation and biodegradation. There were no detections in the water column for any pesticides, which suggests that currently there is no transport of organochlorinated pesticides into Bayou Grande from surface sources.

We tested for 10 trace metals in surface sediments from 78 sites. Selenium was detected in six samples, Cd in 32 samples, As in 53, Hg in 49, Sn in 54, Ni in 71, and Cr, Cu, Pb, and Zn were detected in all 78 samples. The respective TELs were exceeded by As, Cr, Cu, Hg and Ni; the PEL was exceeded by Cd, Pb, and Zn. Selenium and Sn do not have sediment quality guidelines. Because the metals exceed their TEL or PEL they can be assumed to have negative impacts on biota in the Bayou, but their concentrations were generally lower than in Bayous Chico and Texar, two other Pensacola bayous previously studied by PERCH. Concentrations for Hg, and to some extent Ni, are highest in the channel of the main body of the Bayou but most other metals are highest in the upper reaches of the Yacht Basin and Woolsey Bayou near NAS, and in Navy Point Bayou in Warrington. Results for the vibracore samples are consistent with those for the surface samples and show higher levels of trace metals in embayments, especially on the south side of the Bayou. At depth the concentrations are generally lower for the metals that are of anthropogenic origin, but occasionally the TEL is exceeded even at depth.

Principal Component Analysis of the trace metal concentrations indicates that Ni and As are related to the composition of the sediment, and thus are not solely of anthropogenic origin. Another PERCH task, examining trace metal pollution of surface soils, has also found a close relationship with parent material composition for Ni content, but not for arsenic content. Arsenic, however, has a relatively high background concentration in the region and its association with sediment composition is not unexpected. Copper, Pb, Cr, and Zn appear to be of anthropogenic origin. An index that combines the standardized concentrations of these latter metals indicates that activities related to NAS are the main source of anthropogenic trace metals in Bayou Grande and that relatively unpolluted sediments are present in the Bayou near Pensacola Bay in the east and near Jones Swamp in the west.

The present study was the first to assess the environmental quality of the sediments for the entire Bayou Grande systematically and with consistent methods. Many of the pollutants examined exceed regulatory guidelines, including PAHs, dioxins/furans, PCBs and trace metals. Even though these pollutants may be unlikely to directly affect humans, because of limited direct contact of people with the sediments of Bayou Grande, they do have the potential to indirectly affect humans. A case in point are the elevated levels of dioxin/furan and PCB TEQ found in some seafood by another PERCH project, and which seem to be related to sediment contamination identified in the present study. Negative effects on the living environment may be also manifested in reduced populations of some biota.

Bayou Grande is the third urban bayou to be studied by PERCH in Pensacola. The three bayous in common are impacted to varying degrees by either former or current facilities undergoing federally mandated cleanup under the body of law commonly know as superfund. They are also impacted by other industries, commercial activities, and residential activities. The close proximity of NAS Pensacola to Bayou Grande has resulted in impacts to Bayou sediments. These impacts are most severe in parts of embayments that are located closest to runoff areas from the more developed regions of the NAS. In the Woolsey Bayou embayment, impact (above TEL) was observed to depths of 3 meters for PCBs. What is new in the present assessment of Bayou Grande is that there is evidence showing that the Warrington area is also impacting the Bayou. The precise nature of the impacts from both sources is not identical. For example, PAH analyses for samples taken near NAS are characterized by higher concentrations of naphthalenes than those adjacent to Warrington. Relative to the future there is some evidence that suggests that natural degradation is occurring for dioxins/furans and the more highly chlorinated PCB congeners in the deeper sediments. However, other explanations are also possible for the observed temporal changes. In any case, as sediment depth increases, total chlorination declines for dioxins/furans and PCBs. The more chlorinated persistent organic pollutants (POPs) at the surface are not as subject to anaerobic degradation and would be expected to be available to biota. Many sediment bound trace metals such as lead will likely remain bound in the sediment since degradation does not occur. It is possible that POPs and metals will be covered by sediments in the future that will hopefully be less contaminated. The most acute public health issues are related to bacteria contamination originating from fecal sources that poses health risks to recreational users of the bayou’s waters, as described in a previous study. This bacterial contamination can be remediated through the installation of appropriate sewage systems.

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I INTRODUCTION

Bayou Grande, the third urban bayou to be studied by the PERCH Project, is the largest urban bayou in the Pensacola metropolitan area. Through much of its modern history, Bayou Grande has been impacted by avionic support activities and naval vessels at NAS Pensacola. These impacts have elicited environmental and human health concerns. NAS Pensacola has conducted detailed studies of its facility and immediate adjacent waters, including much of Bayou Grande’s southern shoreline. There has only been limited studies of the bayou’s basin or northern shore. The middle and eastern portions of the northern shore are urbanized with residences and commercial activities. The presence of coliform bacteria in Bayou Grande likely derives from these non-military activities. The Escambia County Department of Health has issued frequent health advisories relating to coliform bacteria contamination after rains at Navy Point in Bayou Grande. While it can be assumed that any urbanized waterway will be impacted by anthropogenic activities, there has been no systematic effort to conduct a study of the entire Bayou to verify the presence, magnitude, and origin of substances of concern (SOCs).

The EPA and others estimate that approximately 10 percent of the sediments underlying our nation’s surface waters are sufficiently contaminated with toxic pollutants to pose potential risks to fish and to the humans and wildlife that consume fish and shellfish (USEPA, 1998). Contaminated sediments can affect fish and wildlife by contributing to the bioaccumulation of contaminants in the food chain. The contaminated sediments pose a threat to human health when the pollutants in the sediments bioaccumulate in aquatic organisms routinely consumed by humans. There are numerous examples of cases where fish consumption advisories or bans have been issued for Persistent Organic Pollutants (POPs) such as polychlorinated biphenyls (PCBs), mercury, and dioxins/furans because of the transfer of the pollutants into the food chain (USEPA, 1998). A related PERCH study is presently investigating the bioaccumulation of POPs and other SOCs in seafood tissues (PERCH Task A: Bioaccumulation of chemical contaminants in seafood in the Pensacola Bay region). The objectives of the current investigation are to provide sediment data on POPs such as dioxins/furans and dioxin-like PCBs in support of PERCH Task A, and to conduct analysis of other selected pollutants in Bayou Grande.

II STUDY AREA

II.1 Physiography

Bayou Grande is located in the Warrington area of Pensacola in the southwestern portion of Escambia County in the Florida Panhandle (Figure 1). It is an estuarine water body that has a total surface area of approximately 1.5 square miles (950 acres), a watershed of 10,941 acres (Hatch Mott MacDonald, 2004), and approximately 20 miles of total coastline (EnSafe, 2004). It is the largest of the three urban bayous in Pensacola and is part of the Pensacola Bay System. It still contains substantial undeveloped shoreline on its most western reaches and extends inland for approximately 5 miles into the southwestern portion of Escambia County. Its northern shoreline is bounded by civilian residences and much (approximately 8.5 miles) of the southern shore borders Naval Air Station, Pensacola (NAS) (EnSafe, 2004). The NAS property is 5,800 acres and is located approximately 5 miles southwest of the city of Pensacola. It is surrounded by water on three sides, Bayou Grande to the north, Pensacola Bay to the east, and Big Lagoon and Pensacola Bay to the south (Figure 1) (NAS Pensacola, 2001; Tetra Tech, 2003).

There are extensive wetlands in the southwestern portion of Escambia County with some of these occurring within the watershed of Bayou Grande. These wetlands occur in an area roughly between the NAS and Perdido Bay (Figure 2). While close to Bayou Grande, the Jones Swamp drainage is separate and is part of the Warrington drainage that goes to Bayou Chico to the east. In fact, the drainage of the Bayou Grande area is quite complex, in part because of its young age and minimal relative relief, with some areas of the cited wetlands appearing to drain to both Perdido Bay and Big Lagoon. There is a perception that the area should be largely preserved because of its wetlands and the endangered species which live in it. This area has been of interest to real estate development since it is one of the last places in Escambia County, FL, where waterfront lots and/or large home sites are still available close to the City of Pensacola and Pensacola Naval Air Station. A major controversy over protection of some 7,000 acres to preserve the listed White Tipped Pitcher Plant has occupied a great deal of news media coverage, while land prices have risen as developments occupy more and more of the area. (Droubay et al., 1999). Protection of these areas from human encroachment has also been driven by the need to place protective buffers around NAS and its satellite installations (Droubay et al., 1999).

Bayou Grande’s surrounding watershed is a most likely source for pollutants although some materials may be depositing from the atmosphere either directly into the Bayou or into its watershed. Bayou Grande at 950 acres is the largest Bayou in Pensacola but its watershed of 10,941 acres is small proportionately to the bayou’s total surface. Bayou Texar in east Pensacola, for example, has a surface area of 388 acres (Mohrherr et al., 2005) and yet has a drainage basin almost the size of Bayou Grande’s (10,479 acres). Bayou Chico at 216 acres (Mohrherr et al., 2006) has a drainage basin that is just a little smaller (9,339 acres) (Hatch Mott MacDonald, 2004). This proportionally small drainage area for Bayou Grande implies that, with all other factors being identical, pollutant loads contributed by the drainage should be lower. Airfields such as those at NAS have been shown to produce pollutants, and activities at NAS may contribute to pollution in the Bayou. It is also possible that some pollutants come from Pensacola

Bay and that currents transport them into the Bayou, as has apparently occurred in the past (Collard, 1991).

Figure 1. Location of Bayou Grande.

II.2 Climate

The weather of Bayou Grande is wet, humid, and subtropical with an average annual temperature ranging from 50.5° Fahrenheit in the winter to 82° Fahrenheit in the summer. The average rainfall for the area is among the highest for metropolitan areas in the United States and amounts to approximately 60 inches per year, with the highest amount of rain falling in July and August. Moderate winds tend to prevail from the north during the winter and from the south during the summer (EnSafe, 1999). The area, as in the rest of the northern Gulf of Mexico, is frequently in the direct path of tropical storms and hurricanes. Some abatement of winds and flooding is afforded by Santa Rosa Island and Perdido Key. However, flooding and high wind velocities have caused severe damage during hurricanes to vulnerable structures and boats along the shoreline of Bayou Grande. In September 2004, Hurricane Ivan made landfall as a Category III hurricane about 30 miles west of Bayou Grande, and inflicted heavy damage to the structures in

the area, including those at NAS. Hurricane Ivan’s impact upon the sediments of Bayou Grande is unknown relative to sediment displacement and overall bathymetry.

II.3 Urban Development

The western and northwestern regions of the Bayou are the least developed (Hatch Mott MacDonald, 2004). The eastern half of the southern shore has the NAS airfields, base buildings, and other developed areas. The remaining southern shore is not developed. The eastern half of the northern shore is heavily urbanized with the most densely urbanized area of the watershed being located along Navy Blvd and extending to the west along both sides (north & south) of Gulf Beach Highway until Waycross Street. Past Waycross Street most of the urbanization is to the south of Gulf Beach Highway. A significant portion of the runoff from the most heavily urbanized areas goes into Navy Point Bayou embayment. This embayment has had very high fecal bacterial counts (Snyder, 2006).

Figure 2. Bayou Grande and nearby drainage basins. The map shows that the natural drainage of the southern half of NAS does not flow towards Bayou Grande.

II.4 Pensacola Bay and Dredging Activity

It is likely that the tidal movements into and out of Pensacola Bay have been affected by dredging needed to improve shipping access to Pensacola Bay and estuaries such as Bayou Grande and Bayou Chico. Presently, the Pensacola Harbor navigation channel is used by ships of the US Navy and by commercial vessels going to the Port of Pensacola (Figure 3). The history of dredging goes back to earliest years of United States possession of Pensacola Bay. In 1825, Congress authorized the construction of a navy yard at Pensacola. Congress took this action despite being informed by the Board of Navy Commissioners a year earlier that the channel into Pensacola Bay did not provide at all times a sufficient depth of water for vessels larger than frigates of the first class due to the 21 feet depth of the pass (Pearce, 1989). As early as 1829, it became obvious to some observers that to insure future development at the yard, dredging operations to deepen the channel from the limiting 21 feet identified in the earlier French survey over the bar was needed. Over the years, dredging has deepened the entrance channel into the bay. Currently US Navy Vessels require channel depths in excess of the authorized dimensions of the Civil Works channel. Accordingly, the US Navy has funded the construction and maintenance of the Entrance and Navy Channels, which are currently maintained to the -44 and - 42 foot Mean Lower Low Water (MLLW) depths, respectively. The Civil Works channels within Pensacola Bay have historically been maintained by the US Army Corps of Engineers to a depth of -33 feet MLLW (USEPA and US Army Corps of Engineers, 2005). It is likely that tides and currents have been altered and drastic changes in salinity have taken place due to the dredging but the impact on Bayou Grande or other components of the Pensacola Bay System is unknown. There are presently issues with beach erosion but the specific environmental effects of the dredging have not been studied thoroughly.

Figure 3. Dredged shipping channels in Pensacola Bay.

II.5 Historical Outline of Bayou Grande

The first known documented contact of European colonists with Bayou Grande was in 1559 when the expedition by Don Tristan de Luna settled near what is now the Pensacola NAS. The settlement failed when it lost vital supplies after a hurricane sunk supply boats, and the survivors were not rescued until 1561. The next colonization attempt in 1698 led by Don Andres de Arriola constructed the first permanent post, Fort San Carlos, on the same site de Luna had chosen 140 years earlier. In 1719, war broke out between France and Spain. The French captured the settlement and remained in control for three years. They burned the settlement upon their retreat in 1722.

It was at this time that Pensacola's historic claim of having the finest natural deep-water harbor on the Gulf Coast was made after surveys of it dating from at least 1719. In that year, the master of the Marine Academy of Toulon, France, found that it was a harbor with a bar that would admit ships with a draft of up to 21 feet. Subsequent surveys of the harbor by British cartographer George Gauld in 1764, and by Maj. James Kearney of the U.S. Topographical Engineers in 1822 corroborated the French survey of 1719 (Pearce, 1989). Following the French occupation was a Second Spanish period (1722-1763). The Spanish relocated the settlement to Santa Rosa Island because of the superior defense posture, but hurricanes destroyed the colony. The British controlled West Florida from 1763-1781 then the Spanish again claimed Pensacola for the Third Spanish period (1781-1819) which was contested by Anglo-American settlers and American troops under General Jackson (Pearce, 1989).

In 1821, Pensacola was transferred to the United States. US Naval operations began on Pensacola Bay in 1825, when President John Quincy Adams and Secretary of the Navy, Samuel Southard, established “one of the best equipped naval stations in the country” (NAS Pensacola, 2001). As operations expanded between 1828 and 1835, the Navy acquired approximately 2,300 acres. Figure 4 depicts NAS as it was in about 1860 showing a detailed view of Ft. San Carlos de Barrancas and Ft. Pickens. It also shows the topography of the coast, the U.S. Navy Yard, and all other fortifications from the latest Government surveys in the 1860s (US Department of Defense, 1861). Warrington at that time was located contiguous to the NAS and not at its present location north of Bayou Grande. With the outbreak of hostilities in World War I in Europe, the Navy in 1914 expanded its role in Pensacola’s development by establishing the U.S. Naval Aeronautical Station. As the Navy expanded, so did support businesses and services, bringing more jobs and workers to the area. The military installations have been a major force in Pensacola's growth and development (Global Security, 2007). NAS has over its many years been the site of carrier based aircraft training and maintenance.

Figure 4. Map of Pensacola Bay, 1860s (Exploring Florida, 2008a).

III POLLUTION IN BAYOU GRANDE: LITERATURE REVIEW

III.1 General Pollution History

Prior to 1860 there does not appear to have been any significant urbanization on the northern shore of Bayou Grande on the basis of a map drawn in 1860 (Figure 5) (US Department of Defense, 1861). Logging activities would have certainly occurred due to the ready water access for transport. Such activities would have resulted in some erosion. In Figure 4 Bayou Grande’s connection with Pensacola Bay appears to be mostly fronted by a sand spit that likely severely curtailed tidal exchange with Pensacola Bay except during high wind driven tides and storms. A bridge is seen crossing the Bayou in much the same position as the Navy Boulevard Bridge currently does. While not depicted on the map it is likely that there was some human habitation on the north shore along the road leading to the bridge. The naval base is shown to be situated to the south where a corner of the coast projects into Pensacola Bay. With the outbreak of hostilities in Europe related to World War I more activity occurred at NAS (Global Security, 2007).

Figure 5. Sketch of Pensacola Navy Yard and part of Fort Pickens, 1860s (Exploring Florida, 2008b).

The literature for water and sediment quality in Bayou Grande is not extensive. It is likely that there was significant deterioration of water and sediment quality within the Pensacola Bay System for many years prior to it being “discovered” during the 1950’s. Bayou Grande appears

to have been less impacted by urbanization than is the case with the other major bayous of Pensacola. During the 1950’s when there was a general perception of environmental problems in the Pensacola Bay System, State of Florida scientists were sent to investigate environmental and health conditions in the lower Escambia River, upper Escambia Bay, Bayou Texar, and Bayou Chico (De Sylva, 1955; Murdock, 1955), but apparently not Bayou Grande. Collard (1991) in his review of the Pensacola Bay System did discuss previous studies and reports on Bayou Grande. Butler (1954) had reported that pollution in Pensacola Bay had been a problem for 50 years. Collard (1991) stated that Main Street Sewage Treatment Plant discharges (in operation since 1937) were transported into Bayous Chico and Grande. Portions of Pensacola Bay and Bayous Texar, Chico, and Grande were reported to be of poor bathing quality. Tisdale (1969), as quoted by Collard (1991), reported that NAS discharged significant amounts of oil and grease into Pensacola Bay, but suggested that tidal exchange (flushing) protected Pensacola Bay proper from excessive stress. Water quality measurements showed elevated levels of TKN (Total Kjeldahl Nitrogen), TPO-4(Total Phosphate), BOD (Biochemical Oxygen Demand), and chlorophyll-a west of the Main Street Sewage Treatment Plant and elevated levels of BOD at the mouth of Bayou Grande.

Since Congress passed the Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA) in 1980, the Navy has actively investigated potential contamination that may have resulted from former practices at their installations. In 1988 an environmental permit was issued to NAS under the Resource Conservation and Recovery Act (RCRA). This is a standard permit issued for industrial activities that ensures that ongoing activities are environmentally sound and that spills or leaks of a hazardous waste are investigated and cleaned up (NAS Pensacola, 2003). NAS was placed on USEPA’s National Priority List (NPL) in December 1989. To identify and control environmental contamination, the Navy established the Navy Assessment and Control of Installation Pollutants, which later became part of the Navy’s Installation Restoration Program (IRP). Through these programs, 46 sites at NAS were identified as potential sources of contamination (Naval Facilities Engineering Command, 2004). Some sites are now inactive and are largely without records. Solid wastes have been disposed of primarily at two landfill areas, one site is west of the golf course at the station and the other site is north of Chevalier Field, an airfield. The Bayou Grande Shore Line was identified as Site 40 (Figure 6).

Beginning in the 1930s, industrial wastes from operations at NAS were discharged directly into Pensacola Bay and Bayou Grande. This continued until 1973 when an industrial waste treatment plant began operation. Wastes included paint, solvents, mercury, radium paint, and concentrated plating wastes containing cadmium, chromium, cyanide, lead, and nickel. The plating wastes were discharged via a drainage ditch to Bayou Grande. Other areas of concern include landfills; materials disposal and storage areas; pesticide storage, handling, and disposal areas; solvent, fuel, and industrial waste pipeline leak and spill areas; radium spill areas; and fire and crash training areas (Ecology and Environment, Inc. 1989). Other activities involving hazardous substances include pesticide application, transformer storage and PCBs (USEPA, 2007). Most of these releases could have impacted Bayou Grande either through run-off, aquifer contamination, or atmospheric transport processes. The documented releases are listed and discussed in USEPA (2007) and ATSDR (2006). Other activities that can cause environmental impacts at the NAS are various housing, training, and support activities, as well as the Naval Air Rework Facility

(NARF), a large industrial complex for the repair and overhaul of aircraft engines and frames; the Naval Aviation Depot, which maintains and rebuilds aircraft; and the Navy Public Works Center Pensacola, which provides overall operational support for NAS. Most industrial operations have been conducted in the older portion of the base (Tetra Tech, 2003).

A review of onsite releases was made by the Agency for Toxic Substances and Disease Registry (ATSDR, 2006). The ATSDR, based in Atlanta, Georgia, is a federal public health agency of the U.S. Department of Health and Human Services. The ATSDR is required by law to conduct a public health assessment at each of the sites on the National Priorities List. As part of the public health assessment process, ATSDR conducted a site visit to NAS in February 1991. The visit’s purpose was to collect information necessary to rank the site according to the potential public health hazard it represented and to identify public health issues related to environmental contamination. During the visit, ATSDR staff met base representatives, toured the installation and surrounding areas, and collected community health concerns. Through the Installation Restoration Program, the Navy identified the previously mentioned 46 sites (Figure 6) as potential sources of contamination at NAS Pensacola. ATSDR evaluated the potential for exposure to occur at each of these sites, and identified the following potential exposure situations: • Surface water in Pensacola Bay and Bayou Grande, • Sediments in Pensacola Bay and Bayou Grande, • Fish in Bayou Grande • Blue crabs in Pensacola Bay and Bayou Grande.

The NAS was required by the USEPA to conduct studies that have shown that surface water runoff and groundwater are potential pathways for transport of contaminants to Pensacola Bay, Bayou Grande, and the coastal wetlands (Figure 6). The pollutants present in Bayou Grande originating from NAS fall into five basic categories (EnSafe 1995, 1997, 1998; ATSDR, 2006): 1. Inorganics, common metals and cyanide. 2. Volatile organic compounds originating from solvent used in industrial operations such as electroplating and paint stripping. 3. Semivolatile organic compounds resulting from fuel spills, asphalt, coal, and combustion. 4. Pesticides 5. PCBs

Researchers at the Gulf Ecology Division of USEPA and others also initiated research upon the Bayous of the Pensacola Bay System that included a limited number of samples in Bayou Grande (Lewis et al., 2001). The available data from the non-NAS investigators were compiled into a GIS database by researchers at the Northwest Florida Water Management district (DeBusk et al., 2002). The resulting database incorporated data from FDEP, EPA, NOAA, EMAP, and EPA investigations of the 1980’s and 1990’s (Table 1). The database includes three broad categories of contaminants: heavy (trace) metals, trace organic compounds including PAHs, pesticides and PCBs, and nutrients.

Figure 6. Location of Assessment Zones at NAS. Blue dots represent sampling sites (from: EnSafe, 2004).

Table 1. Sources of data in environmental GIS database for Pensacola Bay System (DeBusk et al., 2002). • George: Master’s Thesis by S. George (1988; University of Southern Mississippi). Data from a comprehensive survey of physical sediment properties in the Pensacola Bay System includes analyses of sediment organic matter. Sampling sites were digitized from maps contained in the printed document. • Sediment Atlas: The raw data set used for the Pensacola Bay system in the Florida Coastal Sediment Contaminants Atlas (FDEP, 1994), obtained from T. Seal via electronic database. Sampling was conducted during the period 1982 through 1991, primarily in Pensacola Bay, upper Escambia Bay, and Bayous Chico and Grande. Site coordinates were provided with the data set. • EPA-Bayous: Sediment studies in bayous of the PBS, conducted by USEPA (Gulf Breeze, FL), Dr. Mike Lewis, principal investigator. Data sets comprising both published and unpublished data were obtained from studies in Bayous Texar and Chico (1993-1994). Site coordinates were provided with the data sets. • NOAA: Sediment analytical data for Pensacola Bay were transcribed from the NOAA (1997) report entitled “Magnitude and Extent of Sediment Toxicity in Four Bays of the Florida Panhandle: Pensacola, Choctawhatchee, St. Andrew and Apalachicola”. Sampling was conducted during 1993-1994 at several sites throughout the Pensacola Bay System. Site coordinates were provided with the data set. • EMAP; EPA-92; EPA-96: Three databases, in electronic format, were obtained from USEPA Region 4 offices, containing results from 1) the 1991-94 EMAP sampling, 2) the 1992 Pensacola Bay Intensive Study and 3) the 1996 Pensacola Bay Study, all conducted by the EPA Gulf Ecology Division at Gulf Breeze. Sampling stations were located throughout the Pensacola Bay System. Site coordinates were provided with the data set.

The NAS sponsored studies of Bayou Grande conducted in the 1990’s have only concerned the areas of Bayou Grande that border the NAS property. These areas have been divided into four assessment zones (AZ): Assessment Zones 1, 2, 3, and 4 as illustrated in Figure 6 (EnSafe, 2003).

The AZ-1 includes portions of the NAS Pensacola shoreline along Bayou Grande from a point near Soldiers Creek to Deepwater Point. Sediments within this zone are mostly fine-grained and characteristic of a low-energy tidal regime. Very few contaminant source areas were identified for this AZ. Potential sources include installation restoration program (IRP) Site 3 and Forrest Sherman Field, which lie south of the zone.

AZ-2 extends from Deepwater Point to J. Kee Point and includes Redoubt Bayou. The shoreline in this area is characterized by sandy beaches with shallow, broad, sandy shelves extending out into the bayou in some areas. In these areas, fine-grained sediment is found further offshore than in AZ-1. The major contributing source to this area is IRP Site 1, potentially contributing inorganics (metals), volatile organic compounds (VOCs), semivolatile organic compounds (SVOCs), and pesticides. Wetlands that surround Site 1 discharge into this zone. A wetland known as the Southeast Drainage Ditch, conveys storm water from the eastern end of Forrest

Sherman Field to the southern end of Redoubt Bayou. This wetland is intersected by an unnamed drainage ditch which passes the south side of Site 16, and conveys surface water from the Barrancas Cemetery area. This intersecting ditch also receives stormwater from an outfall draining the NAS Public Works Center (encompassing IRP Sites 8, 17, 22, and 24). Other wetlands that discharge into the zone include Wetlands 19, 22, 24, and 68 (Figure 6). Contaminants have been detected in some monitoring wells near the shore of this AZ (Tetra Tech, 2003)

AZ-3 extends from J. Kee Point to the Navy Boulevard Bridge. Sediments in this zone are similar to those in AZ-2, with areas of sandy bottom parallel to the shoreline or extending into the bayou as sandbars. Primarily, pesticides from the NAS Pensacola Golf Course may be expected in this area. Contaminants may have been transported to this zone from Site 1 through Wetlands 3 and 4 (Figure 6). A skeet shooting range was located on the east side of Site 1. Wetland 65 also discharges into this zone.

AZ-4 extends from the Navy Boulevard Bridge to the pass connecting Bayou Grande with Pensacola Bay. This area includes Woolsey Bayou, Navy Yacht Basin (Buddy’s Bayou) and portions of Bayou Grande just north of the Navy Yacht Basin. Sediments in this zone are similar to those in AZ-3, with small areas of sandy bottom along the shore. A railroad bridge was formerly in the area.

III.2 Fecal Coliform Pollution in Bayou Grande

Snyder (2006) reported on chronic fecal contamination of waterways in the Pensacola Bay system and found that fecal bacteria presented a public health and environmental problem. The study was a multiyear project, carried out as part of CEDB’s activities, to identify sources of loadings of fecal contamination within the urban bayous of Pensacola. Thirty-one stations were established along the shoreline of Bayou Grande. These stations were selected to coincide with storm water drainages, perennial streams, and areas of likely groundwater discharge indicated by topography and freshwater wetland plants in salt water areas. Samples were taken at monthly intervals from December 1999 to October 2001. Bacterial counts were mostly within the acceptable range but changed drastically with rainfall (Figure 7). Rainfall tends to flush fecal bacteria out of feeder streams and other sources into Bayou Grande and often result in health advisories. The residential areas of the northern and western drainages, and not the Naval Air Station along the southern shore, appeared to be the major source areas for chronic fecal contamination. GIS analyses indicate that older residential developments using septic tanks in low-lying areas are the main source areas (Snyder, 2006). The Escambia County Health Department of Florida continues to monitor Bayou Grande.

Figure 7. Geomeans of Enterococcus at zero rainfall for Bayou Grande.

Concentrations of 35 or lesscolony forming units are considered not to pose a risk to using the waters for recreational activities. This concentration is based on a criterion from the USEPA that has officially announced a final rule for Enterococci criteria for Florida’s Coastal Recreational Waters (marine coastal waters including estuaries). This rule provides a 30-day geometric mean of 35 colony-forming units per 100 milliliters (cfu/100 ml) or less to be considered safe for swimming and water contact sports, and a single sample maximum of 104 cfu/100 ml or less at Designated Bathing Beaches (FDEP, 2007).

III.3 Total Petroleum Hydrocarbons

Total petroleum hydrocarbons (TPH) may be present in Bayou Grande with the most likely origin being aviation fuels used at NAS. Common types of military aviation turbine fuels (turbojet or turbo-prop) are identified by grade designations such as JP-4, JP-5, JP-8, etc. They are among the lighter components that can be detected by the FL-PRO that has a quantitative detection for the range of hydrocarbons going from C8-C40. Jet fuels like JP-4 are composed of about 50-60% gasoline and 40-50% kerosene, are highly volatile, and contain hydrocarbons in

the C4-C16 range. JP-4 was the primary fuel of the USAF for decades; and has been phased out in favor of JP-8. Jet fuel JP-5 is a low-volatility (C10-C19 range) jet fuel with a relatively high flash point (for shipboard safety reasons) and is designed for use in aircraft aboard Navy aircraft carriers. JP-8 was developed to be less volatile and explosive than JP-4. It is safer and has less of an environmental impact. It also contains a full military additive package including a corrosion inhibitor, anti-icing, and anti-static compounds. JP-8 is essentially commercial Jet A-1 fuel with the full military additive package. JP-8+100 is an improved JP-8 fuel with additional “fuel injector cleaner”-type additives that improves the thermal stability of JP-8. An additional difference is significantly less benzene in the environment. As seen in Table 2 aromatics by volume percent compose about 25% of all fuels. These fuels if present should be detectable by the FL-PRO that detect ranges of eight to 40 length hydrocarbon chains since aviation fuels do have hydrocarbons that exceed C-8 in length (USAF, 2006).

Table 2. Chemical and physical requirements for JP-4, JP-5, and JP-8. Issuing Agency: USN USN USAF Grade Designation: JP-4 (NATO F-40) JP-5 (NATO F-44) JP-8 (NATO F-34) Fuel type: Wide-cut gasoline Kerosene type Kerosene type

Acidity, Total (mg KOH/g) 0.015 0.015 0.015 Aromatics (vol %) 25 25 25 Sulfur, Mercaptan (wt %) 0.002 0.002 0.002 Sulfur, Total (wt. %) 0.40 0.30 0.30

III.4 Polycyclic Aromatic Hydrocarbons (PAHs)

The PAHs are compounds composed of two or more aromatic (benzene) rings. PAHs may be divided into two groups, depending upon their mass: low-molecular-weight PAHs, containing two or three aromatic rings, and high-molecular-weight PAHs, containing more than three aromatic rings. PAHs can have multiple origins with oil spills and combustion products being important sources in typical urban environments. They are released into the environment by incomplete combustion and pyrolysis of organic materials such as coal, wood, fuel, garbage, tobacco, and meat. A major source of ambient PAHs is believed to be motor vehicle combustion emissions, particularly in urban areas. Motor vehicle emissions can contribute 46-90% of the mass for individual PAHs in ambient airborne particles in urban areas (Dunbar et al., 2001; Harrison et al., 1996; Nielsen, 1996). PAHs are not particularly soluble in water, but adsorb well to particulate matter, and are therefore usually concentrated in soil or attached to dust particles or marine sediments. PAHs are known to cause environmental deterioration upon accumulating in sediments, as reflected in FDEP sediment quality assessment guidelines (Table 3). Removal of PAHs from the environment occurs more rapidly for 2-ring forms than for the heavier forms via volatilization and biodegradation. Under anaerobic conditions the lighter forms will degrade under nitrate and sulfide reducing conditions and the heavier forms (4-6 rings) tend to adsorb to sediments becoming less available than the lighter 2 and 3-ring forms. The 3-ring forms due to their solubility and volatility exert more acute toxic effects (Brenner et al., 2002).

Table 3. PAH SQAGs1 (μg/kg) and IARC listing (MacDonald, 1994). PAH Compound TEL2 ug/kg PEL3 ug/kg Carcinogenic IARC4 LMW (Light Molecular Weight) PAHs Acenaphthene 6.71 88.9 No listing Acenaphthylene 5.87 128 No listing Anthracene 46.9 245 Not classifiable Fluorene 21.2 144 Not classifiable 2-methylnaphthalene 20.2 201 No listing Naphthalene 34.6 391 Possibly Phenanthrene 86.7 544 Not classifiable Sum LMW-PAHs5 312 1,442 No listing HMW (Heavy Molecular Weight) PAHs Benz(a)anthracene 74.8 693 Probably Benzo(a)pyrene 88.8 763 Probably Chrysene 108 846 Not classifiable Dibenzo(a,h)anthracene 6.22 135 Probably Fluoranthene 113 1,494 Not classifiable Pyrene 153 1,398 Not classifiable Sum HMW-PAHs6 655 6676 No listing Sum LMW&HMW7 1684 16,770 No listing PAHs not assigned SQAG by FDEP Benzo(b)fluoranthene na na No listing Benzo(g,h,i)perylene na na No listing Benzo(k)fluoranthene na na No listing Indeno(1,2,3-cd)pyrene na na No listing 1-Methylnaphthalene na na No listing 1: SQAGs: Sediment quality assessment guidelines adopted by the FDEP. 2: TEL: Threshold effects level (McDonald, 1994). Within this range, concentrations of sediment-associated contaminants are not considered to represent significant hazards to aquatic organisms. 3: PEL: Probable effects levels (McDonald, 1994), lower limit of the range of contaminant concentrations that are usually or always associated with adverse biological effects 4 IARC: The International Agency for Research on Cancer is part of the World Health Organization. Agents with sufficient evidence of carcinogenicity in experimental animals and inadequate evidence of carcinogenicity in humans will ordinarily be placed in the category possibly carcinogenic to humans. When there is strong evidence that carcinogenesis in experimental animals is mediated by mechanisms that do operate in humans, the agent may be upgraded to probably carcinogenic to humans. The classification scheme allows for down-grading to not classifiable as to its carcinogenicity to humans if there is strong, consistent evidence that the mechanism of carcinogenicity in experimental animals does not operate in humans or is not predictive of carcinogenic risk to humans. 5: Sum LMW-PAHs refers to FDEP TEL and PEL values determined for the sum of 7 light molecular weight PAHs

6: Sum HMW-PAHs refers to FDEP TEL and PEL values determined for the sum of 6 heavy molecular weight PAHs 7: Sum LMW&HMW refers to the sum of the concentrations of each of the 13 low and high molecular weight PAHs having FDEP SQAG. While the mode of action of LMW and HMW PAHs is thought to differ, these substances are sometimes grouped in assessments of sediment quality. This results in a derivation of a TEL of 1,684 ug/kg and a PEL of 16,770 ug/kg. (MacDonald, 1994). The actual Total PAH from 8270C SIM analyses includes an additional 5 PAHs.

PAHs can be transported to aquatic sediments via groundwater discharge from an aquifer, via stormwater deposition, air deposition from grass and forest fires and vehicle exhaust, and petroleum product spills. PAHs tend to partition from water into sediments at ratios based on their molecular weight. Larger PAHs tend to partition preferentially into sediments with relatively small concentrations showing up in water due to their lower solubilities. After a spill on the ground the heavier of the PAHs upon entering an aquifer sink to the bottom of the aquifer to form a layer of Dense Non-Aqueous Phase Liquid (DNAPL). The lighter PAHs float on water as the Light Non-Aqueous Phase Liquid (LNAPL) layer. Lighter PAHs such as naphthalene also tend to be more soluble and enter the water column and then evaporate to the atmosphere or are transformed to alkyl forms (Van Mouwerik et al., 1998).

The PAH data for Bayou Grande in the DeBusk GIS database have only two concentrations above the FDEP sediment guidelines (Table 4, Figure 8).

Table 4. Existing ‘non-NAS’ Bayou Grande PAH data [ug/kg] (DeBusk et al., 2002). Site Label TOT_LMW_PAH TOT_HMW_PAH LMW&HMW TEL 312 655 1684 PEL 1442 6676 16770 EPA17 203 1014 1218 EPA18 10 18 28 EPA19 69 180 249 PCOLA15 82 396 477 PCOLA16 59 24 83 PCOLA16 27 58 85 PCOLA17 36 57 94 PCOLA17 29 33 63 NOAA1 40 1 5 NOAA2 x 40 40 NOAA3 90 x x LA91SR32 194 1091 1285

NAS studies show that zone AZ-1 had all PAH concentrations within SQAGs while the other zones, especially AZ-3, had a higher percentage of concentrations that exceeded SQAGs (Table 5) (EnSafe 2003; 2004).

Figure 8. Location of sampling sites from DeBusk et al. (2002) GIS database.

Table 5. PAHs in Bayou Grande sediments by zone of the NAS Pensacola shore (EnSafe, 2003; 2004). Congener TEL AZ1 AZ2 AZ3 AZ4 Aver1 Det2 % Exc3 Aver Det % Exc Aver Det % Exc Aver Det % Exc Naphthalene 34.6 69 7/38 8 75.5 2/55 2 29 2/24 4 58 2/24 8 2-Methylnaphthalene 20.2 94 1/38 3 160 1/55 2 ND 42 2/24 8 1-Methylnaphthalene ND4 ND ND ND Acenaphthylene 5.9 ND ND ND 100 1/24 4 Acenaphthene 6.7 ND 33 1/55 2 5016 2/24 8 35 1/24 4 Fluorene 21.2 ND 34 1/55 2 7900 1/24 4 55 1/24 4 Phenanthrene 86.7 ND 104 7/56 7 1862 15/24 29 119 9/24 17 Anthracene 46.9 ND 80 1/55 2 1440 4/24 13 120 1/24 4 Pyrene 153.0 111 3/38 163 17/57 11 4640 21/24 38 176 15/24 21 Chrysene 108.0 ND 116 13/57 11 2451 21/24 46 139 11/24 21 Benz(a)anthracene 74.8 ND 103 13/57 11 2524 20/24 46 120 11/24 21 Benzo(b)fluoranthene 120 7/38 131 21/57 0 1419 22/24 133 15/24 Benzo(k)fluoranthene ND 110 9/57 0 1121 18/23 133 6/24 Benzo(a) pyrene 88.8 85 1/38 111 16/57 12 1426 20/24 46 133 9/24 17 Indeno(1,2,3-cd) 76 1/38 110 7/57 0 675 19/24 139 5/24 pyrene Dibenz(a,h) 6.2 ND 42 1/55 2 58 3/24 13 ND anthracene Fluoranthene 113.0 107.0 8/38 8 147.0 21/57 16 2881.0 21/24 46 185 15/24 21 Benzo(g,h,i) perylene 85 1/38 110 10/57 0 731 19/24 183 5/24 1 Aver is the average of all concentrations in ug/kg. 2 Det is the number of detections/number of samples. 3 % Exc is the percentage of concentrations that exceed FDEP SQAGs. 4 NDindicates all samples were nondetect and no information was given on the number of samples taken.

III.5 Organochlorinated Compounds

Organochlorinated compounds are diverse groups of compounds that include the organochlorinated pesticides, PCBs, dioxins/furans, PCP, and others. Many organochlorinated molecules and also other halogenated compounds are very persistent both in sediments and when incorporated into living organisms where they can bioaccumulate (USGS, 2005). Use of organochlorinated pesticides and PCBs was widespread beginning in the 1940s until bans and use restrictions were placed on these compounds in the 1970s and 1980s. Compounds such as chlordane, DDT and its environmental degradation products DDD and DDE, and PCBs have low solubilities in water, are strongly associated with organic carbon and fine sediment, and have long environmental half-lives (USGS, 2005). Organochlorinated compounds are included in the persistent organic pollutants (POPs) due to their high recalcitrance to the natural degradative processes.

Pesticides have been applied in the Bayou Grande watershed by NAS base operations (Ensafe, 2003; 2004) and by businesses, local governments and residents for termite and pest control. Many of the commonly applied pesticides are organochlorinated pesticides. The DeBusk et al. (2002) database shows numerous detections for organochlorinated pesticides in Bayou Grande (Table 6, Figure 8). In the NAS studies of pesticides zone AZ-1 had the highest rate of non-detects compared to the other zones (Table 7). Detections of DDT related compounds were generally found in all zones of the NAS shoreline. Zone AZ-2 had the highest number of exceedances of the SQAGs (Ensafe, 2003; 2004).

Table 6. Existing pesticide data for Bayou Grande [ug/kg] (DeBusk et al., 2002). SiteLabel Aldrin A-Chlord G-Chlord 44DDD 24DDD 44DDE 24DDE 44DDT 24DDT TEL None 2.26 2.26 1.22 None 2.07 None 1.19 None PEL None 4.79 4.79 7.81 None 374 None 4.77 None EPA17 1.17 4.62 3.6 0.61 20.18 x 1.73 1.73 EPA18 x 0.14 0.17 x x EPA19 x 0.31 0.67 1.62 0.38 6.37 x x 0.32 PCOLA15 x 0.408 1.477 8.487 7.226 x x x PCOLA15 x x x x 2.938 3.398 x x x PCOLA16 x x x x x x x x x PCOLA16 x x x x x 0.255 x x x PCOLA17 x x x x x 0.477 x x x PCOLA17 x x x x x 0.492 x x x NOAA1 1 1 0.5 x x x x x 0.1 NOAA2 0.5 0.5 1 1 0.5 x x x 0.1 NOAA3 x 4.861 2.58 1 x 2.1 2.02 0.1 LA91SR32 x 0.29 0.44 1.33 0.33 5.1 x 0.37 0.14

Table 6. Existing pesticide data for Bayou Grande [ug/kg] (DeBusk et al., 2002) - continued. SiteLabel Sum_DDT Tot_DDT Dieldr Endo Endo_II Hept Hept_Epox Mirex TEL None None 0.715 None None None None None PEL None None 4.3 None None None None None EPA17 27.85 3.46 x x x 0.11 0.52 x EPA18 0.31 x x 0.15 x x x EPA19 8.69 0.32 x 0.95 x x x PCOLA15 17.19 x 2.019 x x x x 22.156 PCOLA15 6.336 x 0.778 x x x x 14.526 PCOLA16 x x x x x x x 10.395 PCOLA16 0.255 x x 0.35 x x x 10.866 PCOLA17 0.477 x x x x x x 16.624 PCOLA17 0.492 x x x x x x 13.452 NOAA1 7.6 0.6 2.84 x x x x 0.5 NOAA2 3.1 0.6 0.1 x x x x 0.5 NOAA3 7.8 2.12 5.76 x x x x 0.5 LA91SR32 7.27 0.51 x x x x x 1: Italicized underline indicates concentrations above the PEL. 2: Bold type indicates concentrations above the TEL.

For PCBs the DeBusk et al. (2002) database shows that five out of 12 samples were above the TEL, one of which was above the PEL (Table 8). Three Aroclors, Aroclor 1242, 1254, and 1260, were detected in sediments of the NAS portion of the Bayou Grande Shoreline. Aroclor 1242 was only detected in AZ 2, in 8 out 56 samples. Aroclor 1254 was detected in AZ 3 and AZ 4 in 1 out of 24 and 4 out of 23 samples respectively. Aroclor 1260 was detected in all zones in more than 50% of the samples (Ensafe 2003; 2004).

III.6 Trace Metals

Trace metals have been a component of the waste stream generated by NAS (EnSafe, 2003; 2004). Many trace metals are toxic to humans and the environment and SQAGs and water quality standards (Table 9) have been established for trace metals by FDEP. The FDEP listed metals include Arsenic (As), Cadmium (Cd), Chromium (Cr), Copper (Cu), Mercury (Hg), Lead (Pb), Nickel (Ni), and Zinc (Zn). Of these metals mercury is considered to pose the greatest risk to human health.

Data presented in the DeBusk et al. (2002) database for Bayou Grande sediments (Table 10) show that the concentrations of metals were frequently above SQAGs. Arsenic was above the TEL in 8 out 17 samples, cadmium was above the TEL for 9 out of 17 samples with 5 of these exceeding the PEL. Similar results can be observed for copper, lead, mercury, nickel, and zinc. Silver had one sample that exceeded the TEL and tributyltin had no sample that exceeded FDEP guidelines. Overall this data suggests there is definite environmental concern due to the elevated levels of trace metals in Bayou Grande sediments.

Table 7. Pesticides and PCB Aroclors in Bayou Grande sediments by zone of the NAS Pensacola shore (EnSafe, 2003; 2004). Analyte TEL AZ1 AZ2 AZ3 AZ4 Aver.1 Det.2 % Exc.3 Aver. Det. % Exc. Aver. Det. % Exc. Aver. Det. % Exc. Aroclor-1242 21.6 ND 0.82 8/56 2 ND ND Aroclor-1254 21.6 ND ND 5.3 1/24 6.3 4/23 Aroclor-1260 21.6 15.9 28/38 24 31.5 28/56 29 14.4 13/24 13 12.3 13/21 4 4,4'-DDD 1.22 1.6 1/36 3 4.1 18/56 18 0.36 3/24 0.64 3/22 4 4,4'-DDE 2.07 2.4 13/37 18 3.2 25/56 23 1.48 10/24 8 0.98 4/22 4,4'-DDT 1.19 ND 3.1 17/56 9 0.7 10/24 8 0.57 4/22 4 Aldrin 0.95 3/38 0.81 9/56 0.32 1/24 ND 4/22 Dieldrin 0.715 1.2 6/38 16 0.99 18/56 14 26.7 4/24 8 1 4/22 9 Endosulfan I ND 0.12 3/56 ND 5/24 ND 4/22 Endosulfan II 1.3 1/36 0.69 4/56 0.21 1/24 1.2 4/22 Endosulfan ND 0.41 1/56 0.84 5/24 0.73 2/21 sulfate Endrin 3.3 ND 0.41 9/56 1.1 4/24 2.1 4/22 4 Endrin ND 0.83 1/56 ND 0.6 1/21 Alderhyde Endrin Ketone ND 0.76 5/56 1.7 1/24 Heptachlor 0.81 1/36 ND 0.11 1/24 0.11 1/21 Heptachlor 0.48 3/37 1.2 3/56 0.27 5/24 ND epoxide Methoxychlor ND 1.9 ND ND Alpha-BHC 1.3 1/36 0.67 6/56 0.61 5/24 0.61 10/22 Alpha- 0.53 6/36 0.75 12/56 0.63 5/24 0.22 Chlordane beta-BHC ND 0.29 2/56 ND 0.24 Gamma-BHC 0.32 1.1 6/36 16 0.94 15/56 18 0.64 1/24 4 2.9 4/23 13 Gamma- 0.41 1/37 0.82 5/56 0.33 3/24 0.21 3/22 Chlordane 1 Aver is the average of all concentrations in ug/kg. 2 Number of detections/number of samples. 3 % Exc is the percentage of concentrations that exceed FDEP SQAGs.

Table 8. Existing total PCB data for Bayou Grande [ug/kg] (DeBusk et al., 2002) SiteLabel Total_PCB EPA17 249.51 EPA18 2.8 EPA19 48.62 PCOLA15 70.151 PCOLA15 16.648 PCOLA16 1.022 PCOLA17 2.749 PCOLA17 0.592 NOAA1 53.53 NOAA2 16.76 NOAA3 16.76 LA91SR32 53.34 1: Bold type indicates concentrations above the TEL of 21.55 ug/kg. 2: Italicized underline indicates concentrations above the PEL of 188.79 ug/kg.

Table 9. Water quality standards [ug/l] for Class III marine water bodies1. Metal Marine water As (total) <50 Cd <8.8 Cr+6 <50 Cu <3.7 Ni <8.3 Hg <0.025 Pb <8.5 Se <0.07 Zn <86 From: State of Florida 62-302.530, Criteria for Surface Water Quality Classifications.

A total of 143 sediment grabs were taken for trace metal analysis in the NAS related studies. Zone AZ-1 had the highest amounts of arsenic with 39 percent of the samples being above the TEL (Table 11). In Zone AZ-2, 25 percent of the samples were over the TEL and 13 percent were above the TEL in Zones AZ-3 and 4. Cadmium and chromium concentrations showed similar exceedances of the TEL with most occurring in the sediments of the southern shore of the upper Bayou and declining in the eastern part of the Bayou. Lead and mercury concentrations showed a similar but less marked trend being highest in Zone AZ-1 and lower in Zones AZ-3 and 4. There appears to be the possibility of significant impact from metals for much of the southern shore that is contiguous with NAS, especially in the western part with As, Cd, Cr, and Pb exceeding the TEL in 37 to 39% of the samples (Figure 6, Table 11).

Table 10. Existing data for metals in sediments [mg/kg] in Bayou Grande (DeBusk et al., 2002). Site Label As Cd Cr Cu Pb Hg Ni Ag TBT Zn TEL 7.24 0.676 52.3 18.7 30.2 0.13 15.9 0.733 None 124 PEL 41.6 4.21 160 108 112 0.696 42.8 1.77 None 271 EPA17 13.41 6.62 222 55 272 0.297 28 x 0.007 255 EPA18 0.9 0.11 7.6 1.4 3.5 0.009 0.9 x x 14 EPA19 16.8 1.8 169 24.6 79 0.144 22 x 0.003 140 PCOLA15 7.97 4.77 178.06 38.07 128.94 0.17297 14.21 x x 199.21 PCOLA16 0.85 0.34 2.55 x 3.23 0.00267 0.51 x x 4.71 PCOLA17 1.81 0.31 12.63 2.25 8.16 0.02209 2.37 x x 21.6 NOAA1 19.2 4.9 232 48.1 131 0.296 23.4 0.75 x 246 NOAA2 1.3 0.31 10.9 3.1 12.1 0.023 1.4 0.001 x 23.3 NOAA3 21.6 1.21 164 31.2 85 0.179 21.5 0.3 x 170 LA91SR32 12.5 1.41 119 18.3 59.8 0.215 12.6 0.35 0.003 97 BGE-0001 0.18 0.04 10 0.92 2.2 0.01 x 0.01 x 3.1 BGE-0002 9.4 6.7 44 42 120 0.38 x 0.28 x 220 BGE-0003 22 4.2 250 23 84 0.3 x 0.16 x 190 BGE-0004 0.18 0.03 15 1.7 8 0.02 x 0.01 x 12 BGE-0005 0.34 0.05 4.1 1 2.7 0.08 x 0.02 x 5.9 BGE-0006 3.1 0.88 150 3.7 9.1 0.15 x 0.05 x 30 BGE-0007 0.21 0.09 3.3 0.5 2.8 0.1 x 0.01 x 2.2 1: Bold type indicates concentrations above the TEL. 2: Italicized underline indicates concentrations above the PEL.

Table 11. Metals in Bayou Grande sediments by zone of the NAS Pensacola shore (EnSafe, 2003; 2004). Metal TEL AZ-1 AZ-2 AZ-3 AZ-4 Aver1 Det2 %Exc3 Aver Det %Exc Aver Det %Exc Aver Det %Exc Al 8866 38/38 6082 57/57 3171 24/24 2133 24/24 Sb 0.4 6/23 1.1 7/33 0.3 3/24 0.1 1/4 As 7.24 7.4 34/38 39 5.5 45/57 25 2.8 18/24 13 4.2 9/24 13 Ba 6.7 18/38 5.8 45/57 2.9 22/24 2.5 19/24 Be 0.7 16/38 0.7 18/57 0.7 5/24 0.6 4/24 Cd 0.676 1.7 18/38 37 1.9 21/57 28 2.2 7/24 17 1.4 10/24 17 Ca 1298 38/38 952.0 57/57 1892 24/24 1254 24/24 Cr 52.3 54.5 32/38 37 34.0 57/57 25 26.4 24/24 17 21.1 23/24 17 Co 2.5 18/38 1.8 26/57 1.5 9/24 1.2 9/24 Cu 18.7 10.8 36/38 10.1 49/57 7.0 24/24 6.1 20/24 Fe 11055 38/38 8109 57/57 4409 24/24 3178 24/24 Pb 30.2 29.8 36/38 39 31.1 53/57 28 25.8 24/24 25 13.7 22/24 25 Mg 2967.0 38/38 2107 57/57 1222 24/24 974.024/24 Mn 57.0 38/38 42.2 57/57 35.5 24/24 27.4 24/24 Hg 0.13 0.6 7/38 18 0.3 9/57 14 0.2 2/24 8 0.2 2/24 8 Ni 15.9 6.7 24/38 3 7.2 22/57 3 6.6 6/24 3 3.2 8/24 3 K 1329 34/38 862.0 50/57 472.023/24 393.023/24 Se 1.3 18/38 1.2 18/57 1.2 5/24 1.3 4/24 Ag 0.733 0.4 1/38 0.3 1/57 0.4 1/24 0.3 1/24 Na 9197 38/38 6236 57/57 4156 24/24 3797 24/24 Ti 1.7 2/38 0.8 3/57 0.3 5/24 0.5 5/24 V 16.0 38/38 11.6 57/57 5.8 24/24 5.1 20/24 Zn 124 64.7 27/38 13 56.8 38/57 19 25.9 23/24 8 27.7 19/24 8 1 Aver is the average of all concentrations in mg/kg. 2 Det is the number of detections/number of samples. 3 %Exc is the percentage of concentrations that exceed FDEP SQAGs.

IV ENVIRONMENTAL BACKGROUND OF SEDIMENT POLLUTANTS

IV.1 Polychlorinated Biphenyls (PCBs)

PCBs are called persistent organic pollutants (POPs) due to their toxicity, persistence, and biomagnification as they move up through the food chain. PCBs are among the most stable organic compounds known and do accumulate in animal and human tissues. Even though PCBs are no longer commercially produced in the United States, high levels of these chemicals remain in poultry and fish in various parts of the country. PCBs are highly soluble in lipids and are known to biomagnify (concentrate) in human tissues. Because of the persistence of PCBs in environmental media, analyzing for the presence and concentration of PCBs is important in conducting risk assessments (ATSDR, 2000A). Most PCB manufacture and new use was prohibited in the United States in 1978 under TSCA (the Toxic Substances Control Act). TSCA was enacted in 1976 by Congress to give EPA the ability to track the vast number of industrial chemicals currently produced or imported into the United States.

PCBs are a family of chemical compounds formed by the addition of chlorine to biphenyl (Figure 9). A PCB consists of a two-ring structure comprising two 6-carbon benzene rings linked by a single carbon-carbon bond. There are 10 possible substitution positions for chlorine in the two aromatic rings resulting in 209 possible combinations. Molecules with a single chlorine substituent are called "monochlorobiphenyl" (or just "chlorobiphenyl"). Molecules with two chlorines are called "dichlorobiphenyl", and those with three through ten chlorines, in order, are called: "tri", "tetra.", "penta", "hexa", "hepta", "octa", "nona", and "decachlorobiphenyl". Each of the possible chlorine substitutions results in a specific PCB called a congener. The name of a congener specifies the total number of chlorine substituents and the position of each chlorine. For example: 4, 4'-dichlorobiphenyl is a congener comprising the biphenyl structure with two chlorine substituents, one on each of the two carbons at the "4” position of the two rings. Each of the individual congeners has its own unique chemical, physical, and toxicological properties. Each congener has a unique number based upon the International Union of Pure and Applied Chemistry (IUPAC) naming system (ATSDR, 2000A). There is also a BZ nomenclature (Ballschmiter and Zell, 1980) that is presently identical for congener number, but expresses the chemical name in a different way for some of the congeners.

Figure 9. General PCB structure.

PCBs sold in the USA under the trade name Aroclor were sold based upon their overall level of chlorination and were mixtures of multiple PCB congeners. Each Aroclor mixture contained a different blend of individual PCB congeners (Table 12). Aroclor 1254 indicated a PCB mixture with an overall chlorine content of 54%. It is now known that certain congeners in commercial PCB mixtures weather more rapidly than others upon release to the environment, resulting in PCB mixtures in the environment that can be significantly different than the original product. Because traditional laboratory analyses are intended to detect the Aroclor mixtures of PCBs, individual congeners are not routinely reported. These analysis rely upon chromatographic pattern matching between the environmental sample and pure Aroclor mixtures and it is not uncommon for laboratories to list high concentrations of PCBs as ‘non-detected’ for severely weathered environmental samples. This is not because PCBs are absent but rather because the detected pattern no longer resembles the Aroclor mixtures used as the standard of comparison (Schwartz et al., 1987). The traditional Aroclor approach has been supplanted in recent years by analyses for specific PCB congeners with detection limits less than 1 ppt (part per trillion). Our study for Bayou Grande employed EPA method 1668A for the analyses of the 209 PCB congeners (Fikslin and Santoro, 2003).

Table 12. Major PCB congener constituents of five Aroclors [%]. Congener No. Chlorine substitution Aroclor Aroclor Aroclor Aroclor Aroclor (PCB No.) (IUPAC No.) 1016 1242 1248 1254 1260 4 2,2' 4.36 3.99 8 2,4' 10.30 8.97 18 2,5,2' 10.87 9.36 9.95 28 2,4,4' 14.48 13.30 31 2,5,4' 4.72 4.53 9.31 42 2,3,2',4' 7.05 52 2,5,2',5' 4.35 4.08 8.36 53 2,5,2',6' 6.30 70 2,5,3',4' 6.38 4.75 91 2,3,6,2',4' 5.00 99 2,5,2',3',4' 6.10 101 2,4,5,2',5' 6.98 5.04 110 2,3,6,3',4' 8.51 118 2,4,5,3',4' 8.09 138 2,3,4,2',4',5' 5.01 149 2,3,6,2',4',5' 9.52 153 2,4,5,2',4',5' 8.22 180 2,3,4,5,2',4',5' 7.20 185 2,3,4,5,6,2',5' 5.65

IV.2 Organochlorinated Compounds

IV.2.1 General Notions About 1 billion pounds of pesticides are used each year in the United States to control weeds, insects, and other organisms (USGS, 2000; 2005). About 80 percent of this quantity is used in agriculture. Although the use of pesticides has resulted in increased crop production and other benefits, it has raised concerns about potential adverse effects on the environment and human health. Water is one of the primary pathways by which pesticides are transported from their application areas to other parts of the environment (USGS, 2000; 2005). Many pesticides can also be carried via atmospheric means. Due to the hydrophobicity of these compounds they can bind to soil particles that due to stormwater transport become incorporated into the sediments of streams and estuaries.

Sediment serves as a habitat for benthic biota (such as insects and clams, which are commonly consumed by fish), as both a source and a removal mechanism for some contaminants to and from the stream, and as a vehicle for contaminant transport downstream. Aquatic biota also are important in the food web of terrestrial organisms, with some aquatic biota, such as fish, being consumed by people and wildlife. Most chlorinated pesticides are hydrophobic chemicals and have little or no affinity for water; such chemicals have a low solubility in water, a high solubility in lipids (fats), and a strong tendency to sorb to organic material in soil and sediment. Many hydrophobic chemicals also are resistant to degradation, so they persist for a long time in the environment. Historically, the pesticides of primary concern in sediment and aquatic biota have been the organochlorinated insecticides, such as DDT, which were heavily used in agriculture, termite control, and malaria control programs from the mid-1940s to the 1960s or later (USGS, 2000; 2005).

IV.2.2 Organochlorinated Pesticides in Sediments and Biota A large number of pesticides have been detected in stream sediment and aquatic biota in various studies over the last 30 years (USGS, 2000; 2005). The organochlorinated insecticides DDT, chlordane, and dieldrin have been commonly detected in sediment and aquatic biota, even though their agricultural use in the United States was discontinued during the 1970s.

IV.2.2.1 Aldrin and Dieldrin Aldrin and dieldrin are insecticides with similar chemical structures (USEPA, 2003a) (Figure 10 and Figure 11). They are discussed together here because aldrin quickly breaks down to dieldrin in the human organism and in the environment. Pure aldrin and dieldrin are white powders with a mild chemical odor. Neither substance occurs naturally in the environment. The scientific name for aldrin is 1,2,3,4,10,10-hexachloro-1,4,4α,5,8,8α-hexahydro-1,4-endo,exo-5,8- dimethanonaphthalene. The abbreviation for the scientific name of aldrin is HHDN. Technical- grade aldrin contains not less than 85.5% aldrin. The scientific name for dieldrin is 1,2,3,4,10,10- hexachloro-6,7-epoxy-1,4,4α,5,6,7,8,8α-octahydro-1,4-endo,exo-5,8-dimethanonaphthalene. The abbreviation for the scientific name for dieldrin is HEOD. Technical-grade dieldrin contains not less than 85% dieldrin. The trade names used for dieldrin include Alvit, Dieldrix, Octalox, Quintox, and Red Shield (USEPA, 2003a)

Figure 10. Structure of Aldrin.

Figure 11. Structure of Dieldrin.

Aldrin and dieldrin slowly vaporize to the air, but aldrin vaporizes more readily than dieldrin. Aldrin and dieldrin are commonly found in soil, water, or in homes where these compounds were used to kill termites. They are also found in plants and animals near hazardous waste sites (USEPA, 2003a). From the 1950s until 1970, aldrin and dieldrin were widely used pesticides for crops like corn and cotton. Because of concerns about damage to the environment and potentially to human health, EPA banned all uses of aldrin and dieldrin in 1974, except to control termites. In 1987, EPA banned all uses. These are persistent compounds that bind tightly to soil and then slowly vaporize to the atmosphere. Dieldrin in soil and water breaks down very slowly and plants take in and store aldrin and dieldrin from the soil. Aldrin rapidly changes to dieldrin in plants and animals with dieldrin being stored in the adipose tissues.

IV.2.2.2 Endrin Endrin is a solid, white, almost odorless substance that was used as a pesticide to control insects, rodents, and birds ( ATSDR, 1996). Endrin has not been produced or sold for general use in the United States since 1986.

Figure 12. Structure of Endrin.

Endrin is hydrophobic and is found in ground water and surface water at very low levels. It is more likely to sorb onto sediments of rivers, lakes, and other bodies of water. Endrin is generally not found in the air except when it was applied to fields during agricultural applications. The persistence of endrin in the environment depends highly on local conditions. Some estimates indicate that endrin can stay in soil for over 10 years (ATSDR, 1996)). Endrin may also be broken down by exposure to high temperatures (230°C) or light to form primarily endrin ketone and endrin aldehyde. Little is known about the properties of endrin aldehyde or endrin ketone. It is not known what happens to endrin aldehyde or endrin ketone once they are released to the environment; however, the amount of endrin broken down to endrin aldehyde or endrin ketone is very small (less than 5%). Exposure to endrin can cause various harmful effects including death and severe central nervous system injury. Swallowing large amounts of endrin (more than 0.2 mg/kg of body weight) may cause convulsions and kill a human in a few minutes or hours (ATSDR 1996).

IV.2.2.3 Lindane Lindane is a white crystalline organic solid (Figure 13). Most uses being restricted in 1983, lindane is currently used primarily for treating wood-inhabiting beetles and seed treatment. It is also used as a dip for fleas and lice on pets and livestock, for soil treatment, on the foliage of fruit and nut trees, vegetables, timber, ornamentals, and for wood protection. The vapor is colorless and has a slight musty odor when it is present at 12 or more ppm. Lindane has not been produced in the United States since 1976. However, imported lindane is available in the United States for insecticide use as a dust, powder, liquid, or concentrate. It is also available as a prescription medicine (lotion, cream, or shampoo) to treat and/or control scabies (mites) and head lice in humans.

Figure 13. Structure of Lindane (gamma-hexachlorocyclohexane).

IV.2.2.4 Chlordane Chlordane is a human-made chemical that was used as a pesticide in the United States from 1948 to 1988 (Figure 14). It is sometimes referred to by the trade names Octachlor and Velsicol 1068. It is a thick liquid with a color that ranges from colorless to amber, depending on its purity. It may have no smell or a mild, irritating smell. Chlordane is not a single chemical, but is a mixture of many related chemicals, of which about 10 are major components. Chlordane does not dissolve in water. Therefore, before it can be used as a spray, it must be placed in water with emulsifiers (soap like substances), which results in a milky-looking mixture.

Before 1978, chlordane was used as a pesticide on agricultural crops, lawns, and gardens and as a fumigating agent. Because of concerns over cancer risk, evidence of human exposure and build-up in body fat, persistence in the environment, and danger to wildlife, the EPA canceled the use of chlordane on food crops and phased out other aboveground uses over the next 5 years (ATSDR, 1994; 2002). In 1988, when the EPA canceled chlordane’s use for controlling termites, all approved use of chlordane in the United States stopped. Manufacture for export continues.

Figure 14. Structure of Chlordane.

Chlordane may be transported long distances in the atmosphere. The United States appears to be the main source of chlordane in the air over the North Atlantic. The majority of chlordane in the environment probably enters water as runoff from urban and agricultural soils, and is adsorbed to particulates before entering a body of water. The chlordane repartitions in water and volatilizes rapidly near the water surface. The estimated volatilization half-lives of chlordane from a typical pond and lake is <10 days. Nonetheless, monitoring data indicate that sediment concentrations of

chlordane are much higher than the overlying water, suggesting that volatilization from water may not be as fast as predicted.

Chlordane will bioconcentrate in both marine (bioconcentration factor 3,000-12,000) and fresh water (bioconcentration factor 18,500) in rainbow trout. Chlordane is taken up by rooted aquatic vascular plants both from water and from sediment (ATSDR, 1994; 2002).

IV.2.2.5 DDT DDT (1,1,1-trichloro-2,2-bis (p-chlorophenyl)ethane) is a pesticide that was once widely used to control insects on agricultural crops and insects that carry diseases like malaria and typhus. Technical-grade DDT is a mixture of three forms, p,p’-DDT (85%), o,p’-DDT (15%), and o,o’- DDT (trace amounts). All of these are white, crystalline, tasteless, and almost odorless solids (Figure 15). Technical grade DDT may also contain DDE (1,1-dichloro-2,2-bis (p-chlorophenyl) ethylene) and DDD (1,1-dichloro-2,2-bis (p-chlorophenyl) ethane) as contaminants. DDD was also used to kill pests, but to a far lesser extent than DDT. One form of DDD (o,p’-DDD) has been used medically to treat cancer of the adrenal gland. Both DDE and DDD are breakdown products of DDT. DDT does not occur naturally in the environment. After 1972, the use of DDT was no longer permitted in the United States except in cases of a public health emergency. It is, however, still used in some other areas of the world, most notably for controlling malaria. The use of DDD to kill pests has also been banned in the United States (ATSDR, 2002).

In humans, the nervous system would most likely be affected by people consuming food contaminated with large amounts of DDT. People who swallowed large amounts of DDT became excitable and had tremors and seizures. They also experienced sweating, headache, nausea, vomiting, and dizziness. The same type of effects would be expected by breathing DDT particles in the air or by contact of the skin with high amounts of DDT. Tests in laboratory animals confirm the effect of DDT on the nervous system. No effects have been reported in adults given small daily doses of DDT by capsule for 18 months (up to 35 mg per day). People exposed for a long time to small amounts of DDT (less than 20 mg per day), such as people who worked in factories where DDT was made, had some minor changes in the levels of liver enzymes in the blood. A study in humans found that as the DDE levels in the blood of pregnant women increased, the chances of having a pre-term baby also increased. It should be mentioned, however, that the levels of DDE in the blood at which this was noticed were higher than those currently found in women from the general population in the United States, but not higher than those that may be found in women in countries where DDT is still being used (ATSDR, 2002).

Figure 15. Structure of DDT (di-chlorodiphenyltrichloroethane).

Prior to 1973 when it was banned, DDT entered the air, water, and soil during its production and use as an insecticide. DDT is present at many waste sites, including NPL sites. Releases from these sites may continue to contaminate the environment. Most DDT in the environment in the US is a result of past use but DDT still enters the environment because of its current use in other areas of the world. DDE is only found in the environment as a result of contamination or breakdown of DDT. DDT, DDE, and DDD last in the soil for a `very long time, potentially for hundreds of years (ATSDR, 2002). Most DDT breaks down slowly into DDE and DDD, generally by the action of microorganisms. These chemicals stick strongly to soil, and therefore generally remain in the surface layers of soil. Some soil particles with attached DDT, DDE, or DDD may get into rivers and lakes in runoff. Only a very small amount, if any, will seep into the ground and get into groundwater. The length of time that DDT will last in soil depends on many factors including temperature, type of soil, and whether the soil is wet. DDT lasts for a much shorter time in the tropics where the chemical vaporizes faster and where microorganisms degrade it faster. DDT disappears faster when the soil is flooded or wet than when it is dry. In tropical areas, DDT may disappear in less than a year. In temperate areas, half of the ΣDT (Total sum of DDT and its byproducts) initially present usually disappears in about 5 years. However, in some cases, half of the DDT initially present will remain for 20, 30, or more years (ATSDR, 2002).

In surface water, DDT will bind to particles in the water, settle, and be deposited in the sediment. DDT is taken up by small organisms and fish in the water. It accumulates to high levels in fish and marine mammals (such as seals and whales), reaching levels many thousands of times higher than in water. In these animals, the highest levels of DDT are found in their adipose tissue. DDT in soil can also be absorbed by some plants and by the animals or people who eat those crops. DDT released into water adsorbs to particulate matter in the water column and sediment. Sediment is the sink for DDT released into water. There it is available for ingestion by organisms, such as bottom feeders. It was reported that DDT, DDE, and DDD were still bioavailable to aquatic biota in a northern Alabama River 14 years after 432,000 - 8,000,000 kg of DDT was discharged into a river (ATSDR, 2002). DDT, DDE, and DDD are highly lipid soluble. This lipophilic property, combined with an extremely long half-life is responsible for its high bioconcentration in aquatic organisms. The biomagnification of DDT is exemplified by the increase in DDT concentration in organisms representing four trophic levels sampled from a Long Island estuary (ATSDR, 2002). The concentrations in plankton, invertebrates, fish, and fish-eating birds were 0.04, 0.3, 4.1, and 24 mg/kg respectively on a whole body basis (ATSDR, 2002). It was reported that DDE biomagnified 28.7 times in average concentrations from plankton to fish and 21 times from sediment to amphipods in Lake Michigan. In some cases, humans may be the ultimate consumer of these contaminated organisms (ATSDR, 2002).

IV.2.2.6 Mirex and Chlordecone Mirex is no longer made or used in the United States. Mirex was most commonly used in the 1960s and 1970s. Mirex was used as a pesticide to control fire ants mostly in the southeastern part of the United States. It was also used extensively as a flame retardant additive under the trade name Dechlorane in plastics, rubber, paint, paper, and electrical goods from 1959 to 1972 because it does not burn easily (Figure 16a). All registered products containing mirex were banned in the United States between 1977 and 1978 (ATSDR, 1995). Chlordecone is closely related chemically to mirex, a pesticide whose production cease in 1975. The chemical structure of

chlordecone differs from mirex in that the oxygen of the keto group in chlordecone is replaced by two chlorine atoms in mirex (Figure 16b).

It is not known how mirex directly affects the health of people. However, animal studies have shown that eating mirex can cause harmful effects on the stomach, intestines, liver, and kidneys. Eating mirex can also cause harmful effects on the eyes, thyroid, nervous system, and reproductive system. Since these effects occur in animals, they may also occur in people. Exposure to sufficient amounts of mirex may cause cataracts in animals if they are exposed before or soon after birth. It is not known whether human infants may also develop cataracts; it is not likely that mirex will cause cataracts in adults. Short-term, low-level exposure to mirex may harm reproduction and development in rodents. High-level exposures may result in miscarriage (ATSDR, 1995).

It is not known for sure whether mirex causes cancer in humans. Department of Health and Human Services has determined that mirex and chlordecone may reasonably be expected to be carcinogens (ATSDR, 1995). The International Agency for Research on Cancer has determined that mirex and chlordecone are possibly carcinogenic to humans. The EPA has not classified mirex or chlordecone as to carcinogenicity. In rodents, mirex causes liver, adrenal, and blood cancer. Chlordecone also causes liver cancer in rodents, but because of problems with these animal studies more information is necessary to be sure (ATSDR, 1995).

Figure 16a. Structure of Mirex.

Figure 16b. Structure of Chlordecone.

Production of mirex ceased in 1976, at which time industrial releases of this chemical to surface waters were also curtailed. However, releases from waste disposal sites continue to add mirex to the environment. As a pesticide, mirex was widely dispersed throughout the southern United States where it was used in the fire ant eradication program for over 10 years. Adsorption and volatilization are the more important environmental fate processes for mirex, which strongly binds to organic matter in water, sediment, and soil. When bound to organic-rich soil, mirex is highly immobile; however, when adsorbed to particulate matter in water it can be transported great distances before partitioning out to sediment. Atmospheric transport of mirex has been reported based on its detection in remote areas without anthropogenic sources, although this is not a major source of mirex in the environment. Given the lipophilic nature of this compound mirex is both bioaccumulated and biomagnified in aquatic and terrestrial food chains (ATSDR, 1995).

Mirex is a very persistent compound in the environment and is highly resistant to both chemical and biological degradation. The primary process for the degradation of mirex is photolysis in water or on soil surfaces; photomirex is the major transformation product of photolysis. Aerobic biodegradation on soil is a very slow and minor degradation process. Twelve years after the application of mirex to soil, 50% of the mirex and mirex-related compounds remained on the soil (ATSDR, 1995). Between 65-73% of the residues recovered were mirex and 3-6% were chlordecone.

Mirex has been detected at low concentrations in ambient air (mean 0.35 pg/m3) and rainfall samples (<0.5 ng/L) (ATSDR, 1995). In addition, the compound has been detected in drinking water samples. The high bioconcentration factor values (up to 15,000 for rainbow trout) observed for mirex indicate that this compound will be found in high concentrations in aquatic organisms that inhabit areas where the water and sediments are contaminated with mirex. Fish taken from Lake Ontario, the St. Lawrence River, and the southeastern United States (areas where mirex was manufactured or used as a pesticide) had the highest mirex levels (ATSDR, 1995). There are currently eight fish consumption advisories in effect in three states (New York, Pennsylvania, and Ohio) that were triggered by mirex contamination in fish. Waterfowl and game animals have also been found to accumulate mirex in their tissues. Data on mirex residues in foods do not show a consistent trend with regard to contaminant levels or frequency of detection. Mirex has been irregularly detected in Food and Drug Administration (FDA) Pesticide Residue Monitoring Studies since 1978. Little information is available on the specific foods in which residues were found or the levels that were detected. General population exposure to mirex has been determined as a result of several monitoring studies. Levels of mirex in most tissues are very low (at or near the detection limit) (ATSDR, 1995).

Because mirex is a very hydrophobic compound with a low vapor pressure, atmospheric transport is unlikely. When released to surface waters, mirex will bind primarily to the dissolved organic matter in the water with a small amount remaining in the dissolved fraction. Algae are known to bioaccumulate mirex with bioaccumulation factors (BCF) in the range of 3,200-7,300, while bacteria have a BCF of 40,000. Based on a water solubility of 0.6 mg/L, a bioconcentration factor of 820 was calculated for mirex ATSDR, 1995). Bioaccumulation of mirex also occurred in invertebrates exposed to 0.001-2.0 μg/g mirex in water; tissue residues ranged from 1.06 to 92.2 μg/g. A laboratory BCF for mirex in rainbow trout was 1,200 and an actual BCF found in

rainbow trout in Lake Ontario was 15,000. Biomagnification of mirex is supported by a study of various aquatic organisms that comprise an aquatic food chain. In a 1972 residue study conducted in Mississippi during the time when mirex was being used extensively in fire ant control programs, mirex accumulation in grassland invertebrates (e.g., spiders and grasshoppers) ranged from 100 to 700 μg/kg (ppb) (ATSDR, 1995).

IV.2.2.7 Endosulfan Endosulfan is a manufactured pesticide. It is used to control a number of insects on food crops such as grains, tea, fruits, and vegetables and on nonfood crops such as tobacco and cotton. It is also used as a wood preservative. It is a cream to brown colored solid that may appear crystalline or in flakes (Figure 17). It has a distinct odor similar to turpentine. (ATSDR, 2000b).

Symptoms of endosulfan poisoning have been seen in some people who were exposed to very large amounts of this pesticide during its manufacture. Symptoms of endosulfan poisoning have also been seen in people who intentionally or accidentally ate or drank large amounts of endosulfan. Most of these people experienced convulsions or other nervous system effects. Some people who intentionally ate or drank large amounts of endosulfan died. The health effects in people exposed to smaller amounts of endosulfan for longer periods are not known. It is not known whether endosulfan has ever affected the ability of people to fight disease or has ever caused cancer in people. Endosulfan has not been classified as to its ability to cause cancer (ATSDR, 2000b).

Figure 17. Structure of Endosulfan.

Endosulfan released to the atmosphere may be transported for long distances before being removed in wet and dry deposition. Endosulfan sulfate is a reaction product found in technical-grade endosulfan as a result of oxidation, biotransformation, or photolysis. There is very little difference in toxicity between endosulfan and its metabolite, endosulfan sulfate. Results of several laboratory and greenhouse studies indicate that endosulfan is strongly adsorbed to soil. In standard glass-column elution tests, it was found to adsorb tightly to loamy sand, sandy loam, sandy clay loam, and sandy clay soils (ATSDR, 2000b).

Endosulfan does not bioaccumulate to high concentrations in terrestrial or aquatic ecosystems. In aquatic ecosystems, residue levels in fish generally peak within 7 days to 2 weeks of continuous exposure to endosulfan. Maximum bioconcentration factors are usually less than 3,000, and residues are eliminated within 2 weeks of transfer to clean water. A maximum Bioconcentration

Factor (BCF) of 600 was reported for endosulfan in mussel tissue (ATSDR, 2000b). In a similar study, endosulfan had a measured BCF of 22.5 in mussel tissue. Tissue concentrations of endosulfan fell rapidly upon transfer of the organisms to fresh seawater; for example, a depuration half-life (removal half-life) of 34 hours. Higher BCFs were reported for whole-body and edible tissues of striped mullet (maximum BCF=2,755) after 28 days of exposure to endosulfan in seawater. However, tissue concentrations decreased to undetectable levels 48 hours after the organisms were transferred to uncontaminated seawater. Similarly, a BCF of 2,650 was obtained for zebra fish exposed to 0.3 μg/L of endosulfan for 21 days in a flow- through aquarium (ATSDR, 2000b). In freshwater studies, mosquito fish, catfish, and freshwater eels were exposed to endosulfan in static tests. Maximum tissue concentrations in mosquito fish (933 μg/kg) were found in fish exposed to 16 μg technical-grade endosulfan/L for 24 hours. The maximum tissue concentrations in fish exposed to 2 μg technical-grade endosulfan/L for 7 days was 143 μg α-isomer/kg. (ATSDR, 2000b).

Although endosulfan and endosulfan sulfate have been found at low concentrations in a few surface water and groundwater samples collected at hazardous waste sites, no information was found in the available literature regarding current concentrations of endosulfan or endosulfan sulfate in domestic surface waters not associated with these sites. The World Health Organization (WHO) reported that although endosulfan has been detected in agricultural runoff and in surface waters draining industrialized areas, contamination of surface waters with this compound does not appear to be widespread and stated that endosulfan concentrations in surface water are generally <1 ppb. In a survey of streams in the western United States conducted by the U.S. Geological Survey (USGS) from 1968 to 1971, endosulfan was detected in only 1 of the 546 surface water samples collected, at a concentration of 0.02 μg/L (ATSDR, 2000b).

Endosulfan has been detected in only a limited number of urban and agricultural soils in the United States. The National Soils Monitoring Program conducted in 1972 included the collection of 1,483 soil samples from 37 states. Endosulfan and endosulfan sulfate were each detected in only one sample at <0.01 ppm. Endosulfan was not detected (method detection limit of 1 μg/kg) in sediments collected from the Central Columbia plateau of the United States (ATSDR, 2000b).

IV.2.3 Dioxins/Furans Dioxins/furans and dioxin-like compounds are ubiquitous environmental contaminants that are very stable against chemical and microbiological degradation and therefore persistent in the environment. The major sources of dioxins/furans are combustion processes, such as waste incineration and metal smelting and refining. Dioxin/furan contamination is also associated with the production and the use of pentachlorophenol (PCP) at wood treating sites. Other sources of localized dioxin/furan hotspots include spills from PCB filled electrical equipment such as transformers and capacitors, each of which may contain several kilograms of PCBs and hundreds of milligrams of dioxins/furans. Paper mills in the past produced dioxins/furans (mainly the 2,3,7,8-TCDD congener) during a chlorine bleaching process (ATSDR, 1998). Agent orange, a herbicide once employed by the Department of Defense, and tested at Eglin Air Force Base about 60 miles east of Bayou Grande, contained as an active ingredient the herbicide 2,4,5-T that was contaminated with minute amounts of dioxins/furans as a by-product of the manufacturing process (Frumkin, 2003). Dioxins/furans are fat-soluble and thus tend to bioaccumulate in the lipids of animal tissues and in the food chain. Food presumably contaminated by environmental dioxins/furans is the major source for human exposure to dioxins/furans, especially fatty foods:

dairy products (butter, cheese, and fatty milk), meat, eggs, and fish. Some subgroups within the society (e.g., nursing babies and people consuming large quantities of dairy products and fish) may be highly exposed to these compounds and are thus at greater risk (ATSDR, 1998).

Dioxin/furan compounds include polychlorinated dibenzo-p-dioxins (PCDDs) and polychlorinated dibenzofurans (PCDFs). The chlorinated dibenzodioxins include 75 individual compounds and the dibenzofurans include 135 compounds. They are tricyclic aromatic compounds with similar physical and chemical properties. There are also dioxin-like polybrominated dibenzo-p-dioxins (PBDDs or BDDs), polybrominated dibenzofurans (PBDFs or BDFs), polybrominated diphenyl ethers (PBDE) and polychlorinated biphenyls (PCBs). Dioxin- like refers to the fact that some polyhalogenated compounds assume similar structural conformations and similar physicochemical properties that invoke a common battery of toxic responses. The brominated compounds with dioxin-like activity are not included in the current study but the dioxin-like PCBs are.

The most widely studied congener of this general class of compounds is 2,3,7,8- tetrachlorodibenzo-p-dioxin (TCDD). This compound, often called simply “dioxin,” is the most toxic and is used as the reference for calculations of toxicity. The structure of TCDD and several related compounds are shown in Figure 18. These individual compounds are referred to technically as congeners. Out of 75 congeners of dioxins, seven appear to have dioxin-like toxicity. Out of the 135 possible congeners of furans, 10 appear to have dioxin-like toxicity. This makes a total of 17 individual dioxin/furan congeners exhibiting dioxin-like toxicity. Some PCBs are structurally and conformationally similar to dioxins/furans (Figure 18) and 12 of the 209 PCB congeners appear to have dioxin-like toxicity.

Figure 18. Chemical Structure of 2,3,7,8-dioxin (TCDD) and representative dioxin-like compounds.

Dioxins bring about a wide spectrum of biochemical and toxic effects in experimental animals. These effects depend on species, strain, gender, age and tissue. For the most part, the mechanisms of the impacts are still obscure. Dioxins/furans and dioxin-like PCBs persist and bioaccumulate in fatty tissues of animals and humans because of their hydrophobic nature and resistance towards metabolic breakdown in the body. It appears that because they are fat-soluble and not very soluble in water, they cannot be readily excreted in urine and animals are not able to metabolize them. The excretion of these compounds is so slow that their half-life is many years, which means that it takes years for the human body to get rid of 50 % of the compound once it has become incorporated into tissue. Because dioxins/furans are mixtures, every compound has a different half-life, but as a rule of thumb, an average half-life is ten years. This long half-life makes them highly cumulative compounds, i.e., they accumulate in the body over decades even at a low exposure (ATSDR, 1998).

In humans, a wide variety of health effects have been linked to high exposures to dioxins/furans, including mood alterations, reduced cognitive performance, diabetes, changes in white blood cells, dental defects, endometriosis, decreased male/female ratio of births and decreased testosterone and elevated thyroxin levels in neonates. Presently, in humans the effects have been proven to include chloracne (skin disease with severe acne-like pimples). The effect that has caused the greatest public concern is cancer. The USEPA has characterized TCDD as “carcinogenic to humans” (USEPA, 2003b) but there is some uncertainty about the relationship between dioxin and cancer, especially at lower concentrations (National Research Council, 2006). Another health concern are possible developmental effects (ATSDR. 1998).

Dioxin and dioxin-like toxicity is expressed as Toxic Equivalents (TEQ). To determine the TEQ all of the toxic dioxins/furans and dioxin-like PCBs have been assigned a Toxic Equivalency Factor (TEF). TEFs compare the potential toxicity of each dioxin-like compound comprising the mixture to the established toxicity of TCDD, the most toxic dioxin/furan. TEFs were established through review of toxicological databases along with considerations of chemical structure, persistence, and resistance to metabolism. That information has been used to ascribe specific “order of magnitude” TEFs for each dioxin-like congener relative to TCDD, which is assigned a TEF of 1.0. The most recent system of TEQ assessment (Van den Berg et al., 2006) has TEF values for 17 dioxins/furans and 12 PCBs.

IV.2.4 Pentachlorophenol (PCP) Pentachlorophenol (PCP) is a manufactured chemical that does not occur naturally. Technical grade PCP that is used in industry is more toxic than pure PCP due to its by-products. By- product contaminants of PCP production include various dioxins and furans. PCP was widely used as a pesticide and wood preservative but since 1984 the purchase and use of PCP has been restricted to certified applicators. It is no longer available to the general public, but is still used industrially as a wood preservative for utility poles, railroad ties, and wharf pilings. PCP can be found in the air, water, and soil. It enters the environment through evaporation from treated wood surfaces, industrial spills, and disposal at uncontrolled hazardous waste sites. PCP is broken down by sunlight, other chemicals, and microorganisms within a few days to months (Rao, 1978; ATSDR, 2001). Studies of workers show that exposure to high levels of PCP can cause the cells in the body to produce excess heat. When this occurs, a person may experience a very high fever, profuse sweating, and difficulty breathing. The body temperature can increase to dangerous

levels, causing injury to various organs and tissues, and even death. Liver effects and damage to the immune system have also been observed in humans exposed to high levels of PCP for a long time (ATSDR, 2001). The USEPA has determined that PCP is a probable human carcinogen and the International Agency for Cancer Research (IARC) considers it a possible human carcinogen. In drinking water the USEPA and FDEP MCL has been set at 1 ppb or 1 μg/L (ATSDR, 2004). For coastal sediments the AET is 17 ug/kg (NOAA, 1999). A TEL or PEL for coastal sediments has not yet been determined by FDEP.

IV.3 Metals

IV.3.1 Mercury (Hg) Mercury is used in a number of scientific applications and in dental amalgam. Mercury is mostly obtained by reduction from the mineral cinnabar. Mercury occurs in deposits throughout the world and is poisonous in soluble forms such as mercuric chloride or methylmercury. The volatile nature of mercury plays an important role in its environmental fate (ATSDR 1999).

The natural global bio-geochemical cycling of mercury is characterized by degassing of the element from soils and surface waters, followed by atmospheric transport, deposition of mercury back to land and surface waters, and sorption of the compound to soil or sediment particulates. Mercury deposited on land and open water is in part revolatilized back into the atmosphere. This emission, deposition, and revolatilization create difficulties in tracing the movement of mercury to its sources. Particulate-bound mercury can be converted to insoluble mercury sulfide and precipitated or bioconverted into more volatile or soluble forms that re-enter the atmosphere or are bioaccumulated in aquatic and terrestrial food chains (ATSDR 1999).

Mercury has three valence states. The specific state and form in which the compound is found in an environmental medium is dependent upon a number of factors, including the redox potential and pH of the medium. The most reduced form is metallic or elemental mercury, which is a liquid at ambient temperatures, but readily vaporizes. Over 95% of the mercury found in the atmosphere is gaseous mercury (Hg0), the form involved in long-range (global) transport of the element. Although local sources are important, a 72-hour travel time trajectory for mercury indicates that some mercury found in rain may originate from sources up to 2,500 km (1,550 miles) away. Over the last 140 years, the atmospheric mercury concentrations have increased by a factor of 3.7, or approximately 2% per year (STEIN ET AL., 1996). Metallic mercury released in vapor form to the atmosphere can be transported long distances before it is converted to other forms of mercury, and wet and dry deposition processes return it to land and water surfaces. Most inert mercury (Hg+2) in precipitation is bound to aerosol particulates, which are relatively immobile when deposited on soil or water. Mercury is also present in the atmosphere to a limited extent in unidentified soluble forms associated with particulate matter. In addition to wet and dry deposition processes, mercury may also be removed from the atmosphere by sorption of the vapor form to soil or water surfaces (USEPA, 1984).

In soils and surface waters, mercury can exist in the mercuric (Hg+2) and mercurous (Hg+1) states as a number of complex ions with varying water solubilities. Mercuric mercury, present as complexes and chelates with ligands, is probably the predominant form of mercury present in surface waters. The transport and partitioning of mercury in surface waters and soils is

influenced by the particular form of the compound. More than 97% of the dissolved gaseous mercury found in water consists of elemental mercury. Volatile forms (e.g., metallic mercury and dimethylmercury) are expected to evaporate to the atmosphere, whereas solid forms partition to particulates in the soil or water column and are transported downward in the water column to the sediments. The dominant process controlling the distribution of mercury compounds in the environment appears to be the sorption of nonvolatile forms to soil and sediment organic matter. Small amounts (2–4 ng/l) of mercury are able to move from contaminated groundwater into overlying lakes, with concentrations reaching a maximum near the sediment/water interface; however, since most of the mercury in the groundwater is derived from atmospheric sources, this low value indicates that most of the mercury deposited on soil (92 - 96% of the 10.3 μg/m2/year of mercury deposited) is absorbed to the soil and does not leach down into the groundwater (ATSDR 1999).

The sorption process has been found to be related to the organic matter content of the soil or sediment (ATSDR 1999). Mercury is strongly sorbed to humic materials and sesquioxides in soil at a pH higher than 4 and to the surface layer of peat. Mercury has been shown to volatilize from the surface of more acidic soils (i.e., soil pH of less than 3.0). Adsorption of mercury in soil is decreased with increasing pH and/or chloride ion concentrations. Mercury is sorbed to soil with high iron and aluminum content up to a maximum loading capacity of 15 g/kg (15,000 ppm). Inorganic mercury sorbed to particulate material is not readily desorbed. Thus, freshwater and marine sediments are important repositories for inorganic forms of the element, and leaching is a relatively insignificant transport process in soils. However, surface runoff is an important mechanism for moving mercury from soil to water, particularly for soils with high humic content. Mobilization of sorbed mercury from particulates can occur through chemical or biological reduction to elemental mercury and bioconversion to volatile organic forms.

The most common organic form of mercury, methylmercury, is soluble, mobile, and quickly enters the aquatic food chain. This form of mercury is accumulated to a greater extent in biological tissue than are inorganic forms of mercury (Riisgard and Hansen, 1990). Methylmercury in surface waters is rapidly accumulated by aquatic organisms; concentrations in carnivorous fish (e.g., pike, shark, and swordfish) at the top of both freshwater and marine food chains are biomagnified on the order of 10,000–100,000 times the concentrations found in ambient waters. Mercury concentrations in fish have been negatively correlated with water quality factors such as alkalinity and dissolved oxygen content. The biomagnification of methylmercury has been demonstrated by the elevated levels found in piscivorous fish compared with fish at lower levels of the food chain (ATSDR 1999).

Mercury is transformed in the environment by biotic and abiotic oxidation and reduction, bioconversion of inorganic and organic forms, and photolysis of organomercurials. Inorganic mercury can be methylated by microorganisms indigenous to soils, fresh water, and salt water. This process is mediated by various microbial populations under both aerobic and anaerobic conditions. The most probable mechanism for this reaction involves the nonenzymatic methylation of mercuric mercury ions by methylcobalamine compounds produced as a result of bacterial synthesis. Mercury forms stable complexes with organic compounds. Monoalkyl mercury compounds (e.g., methylmercuric chloride) are relatively soluble; however, the solubility of methylmercury is decreased with increasing dissolved organic carbon content,

indicating that it is bound by organic matter in water. Dialkyl mercury compounds (e.g., dimethylmercury) are relatively insoluble. Dimethylmercury is volatile, although it makes up less than 3% of the dissolved gaseous mercury found in water. The major pathways for transformation of mercury and various mercury compounds in air, water, and soil are shown in Figure 19.

Figure 19. Cycling of mercury between the atmosphere, water, sediments, and organisms (Stein et al., 1996).

IV.3.2 Lead (Pb) Native lead occurs in nature, but it is rare. Lead is a ductile, dense metal, resistant to corrosion and a poor electrical conductor. Lead has a long history of use, dating to the earliest civilizations. In fact, some of the ancient uses of lead, such as for plumbing, cosmetic and medicinal purposes, are still used today, albeit under extreme warnings and restrictions. Early uses in the U.S. were primarily for ammunition, brass and pewter, paints and protective coatings, glass and crystal, ceramic glazes, and water lines and pipes. Ultimately, lead use expanded to include machine bearings, cable covering, caulking, solder, fuel additives and lead acid storage batteries (Ostrom et al., 2004).

In the twentieth century, the use of lead paralleled certain technological advances. Developments in applications of electricity and telecommunications resulted in the use of lead for machine bearings, cable covering, caulking and solder. Similarly, as motor vehicles became commonplace, the demand for lead increased due to its use in lead acid storage batteries and fuel additives. In the 1970s, amid growing concern over the health effects associated with lead exposure, environmental regulations restricted the use of lead in certain consumer products and activities in the U.S. Two of the most notable of these are lead-based paint and leaded gasoline (Ostrom et al., 2004).

Lead-based paint was formerly extensively used and being applied as an exposed surface covering was weathered and often carried ultimately by stormwater to estuaries like Bayou Grande. As concern regarding the manufacture and application of lead paint continued to grow, its use was restricted overseas. In 1922, the Third International Labor Conference of the League of Nations recommended banning interior uses of white lead (as lead paint was referred to at that time). By 1934 a number of countries, including France, Belgium, Austria, Tunisia, Greece, Czechoslovakia, Great Britain, Sweden, Poland, Spain, Yugoslavia and Cuba, had restricted the interior use of lead paint. In the U.S., paint manufacturers set a voluntary standard only in 1960, limiting the amount of lead in paint to five percent. In 1971, the Lead-Based Paint Poisoning Prevention Act initiated a national effort to reduce the hazards associated with exposure to lead- based paint (Ostrom et al., 2004; ATSDR, 2005b).

Lead was first added to gasoline in the 1920s to increase the octane rating of gasoline, thereby reducing engine knock. Before the phase out of leaded gasoline in the U.S., approximately 250,000 tons of organic lead was added each year to gasoline produced in this country (Ostrom et al., 2004; ATSDR, 2005b). The USEPA determined in the 1970s that the lead emissions resulting from leaded gasoline posed a significant health risk to urban populations, especially to children. It has been estimated over the life of a vehicle, 75% of the lead consumed with leaded gasoline was emitted as particulate matter in the exhaust (ATSDR, 2005b). Small particles (<0.1 um) could remain airborne for up to 7 to 30 days and travel thousands of miles from the original source, whereas larger particles, which are formed by the agglomeration of smaller particles, spend less time airborne and don’t travel as far. Over time, as concern over the health effects associated with lead began to grow, health and environmental regulations were enacted to restrict the use of lead in certain products and activities in the U.S. In the last twenty-five years, lead- based paint, leaded gasoline, leaded can solder and lead-containing plumbing materials were among the products that were gradually restricted or phased out of use. The use of lead, however, is far from obsolete. An important commodity, lead is now primarily used to manufacture lead acid batteries. Lead also continues to be used in paint, glass, ceramics, pigments, casting metals, metal products, solder, and other minor uses (Ostrom et al., 2004; ATSDR, 2005b).

Lead is one of the trace metals of greatest water-quality concern in urban areas and highway stormwater runoff. Its former use as an additive in gasoline has caused widespread contamination of soils near highways and streets and in drainage ways for stormwater runoff from these areas. Also of concern is its continued presence in gasoline at “natural” concentrations to cause highway and street stormwater runoff from some areas to have lead at sufficient concentrations to violate USEPA water-quality criteria for stormwater runoff (Ostrom et al., 2004; ATSDR, 2005b).

As a result of its extensive use throughout history, lead is ubiquitous in the environment. Manmade sources of lead in the environment far exceed natural sources. In water, elemental lead is insoluble and most other lead compounds have low water solubilities. The occurrence of dissolved lead in surface water depends primarily on the pH and salt content of the water. In theory more dissolved lead can be expected in soft waters with low pH. Calculations show that at pH<5.4 the total lead solubility is about 30 ug/L in hard water and about 500 ug/L in soft water. In water with a pH near 6.5 and an about 25 mg bicarbonate ion/l, common in areas of the northeastern U.S., lead concentrations of 330 ug/l can be stable. In most natural water, however,

lead tends to form compounds of low water solubility with anions in the water, such as hydroxides, carbonates, sulfates and phosphates. These compounds ultimately precipitate out of the water and either adsorb and accumulate in sediments or are incorporated into organic matter and other solid particles carried in the water. In river water most lead is expected to be in an undissolved form (Ostrom et al., 2004; ATSDR, 2005b).

The degree of bioaccumulation in aquatic or terrestrial food chains is not clear. Older organisms tend to contain the greatest body burdens of lead. In aquatic organisms, lead concentrations are usually highest in benthic organisms and algae, and lowest in upper trophic level predators (e.g., carnivorous fish). However, exposure of a fresh-water fish to several sublethal concentrations of lead for a period of 30 days showed significant accumulation of lead in the blood and tissues (ATSDR, 2005b). High BCFs were determined in studies using oysters (6,600 for Crassostrea virginica), fresh-water algae (92,000 for Senenastrum capricornutum), and rainbow trout (726 for Salmo gairdneri). However, most median BCF values for aquatic biota are significantly lower: 42 for fish, 536 for oysters, 500 for insects, 725 for algae, and 2,570 for mussels. Lead is toxic to all aquatic biota, and organisms higher up in the food chain may experience lead poisoning as a result of eating lead-contaminated food (ATSDR, 2005b).

IV.3.3 Cadmium (Cd) Cadmium most often occurs in small quantities associated with zinc ores. Cadmium is a metal with a high toxicity that is not known to be essential for plant or animal life. Cadmium is toxic at very low exposure levels and has acute and chronic effects on health and environment. New releases add to the already existing deposits of cadmium in the environment. Cadmium and cadmium compounds are, compared to other metals, relatively water-soluble. They are therefore also more mobile in soil, more bioavailable, and tend to bioaccumulate (Nordic Council of Ministers, 2003).

Almost all cadmium is obtained as a by-product in the treatment of zinc, copper, and lead ores. It is a component of some of the lowest melting alloys; it is used in bearing alloys with low coefficients of friction and great resistance to fatigue; it is used extensively in electroplating, which accounts for about 60% of its use. Cadmium accumulates in the human body and especially in the kidneys. According to current knowledge kidney damage (renal tubular damage) is probably the critical health effect. Other effects of cadmium exposure are disturbances of calcium metabolism, hypercalciuria and formation of stones in the kidney. High exposure can lead to lung cancer and prostate cancer. Atmospheric deposition seems continuously to cause the content of cadmium in agricultural top soil to increase, which by time will be reflected in an increased human intake by foodstuffs and therefore in an increased human risk of kidney damages and other effects related to cadmium. The presence of cadmium in the soil or on surfaces can result in stormwater transport to estuaries. In the marine environment levels of cadmium may significantly exceed background levels causing a potential for serious effects on marine animals and in particular birds and mammals (Nordic Council of Ministers, 2003).

Cadmium and cadmium compounds are, compared to other heavy metals, relatively water soluble. They are therefore also more mobile in soil, generally more bioavailable and tend to bioaccumulate. Cadmium is readily accumulated by many organisms, particularly by

microorganisms and mollusks where the bioconcentration factors are in the order of thousands. Soil invertebrates also concentrate cadmium markedly. Most organisms show low to moderate concentration factors of less than 100. In microorganisms cadmium is toxic as demonstrated by laboratory experiments (Nordic Council of Ministers, 2003). The main effect is on growth and replication. The most affected soil microorganisms are fungi, some species being eliminated after exposure to cadmium in soil. In aquatic organisms cadmium is most readily absorbed by organisms directly from the water in its free ionic form Cd (II). Effects of long-term exposure can include larval mortality and temporary reduction in growth (Nordic Council of Ministers, 2003). Sublethal effects have been reported on the growth and reproduction of aquatic invertebrates. The toxicity is variable in fish, salmonoids being particularly susceptible to cadmium. Sublethal effects in fish, notably malformation of the spine, have been reported. The most susceptible life-stages are the embryo and early larva, while eggs are the least susceptible. In studies of lake trout exposed to different levels of cadmium, researchers found that cadmium affected foraging behavior, resulting in lower success at catching prey (Nordic Council of Ministers, 2003).

IV.3.4 Arsenic (As) Arsenic is an element that is widely distributed in the Earth’s crust. Elemental arsenic is ordinarily a steel grey metal-like material that occurs naturally. However, arsenic is usually found in the environment combined with other elements such as oxygen, chlorine, and sulfur. Arsenic combined with these elements is called inorganic arsenic. Arsenic combined with carbon and hydrogen is referred to as organic arsenic. Understanding the difference between inorganic and organic arsenic is important because some of the organic forms are less harmful than the inorganic forms (ATSDR, 2005a).

Inorganic arsenic occurs naturally in soil and in many kinds of rock, especially in minerals and ores that contain copper or lead. When these ores are heated in smelters, most of the arsenic goes up the stack and enters the air as a fine dust. Until recently, about 90% of all arsenic utilization in the US was used for the copper chromated arsenic (CCA) production of treated wood. In 2003, U.S. manufacturers of wood preservatives containing arsenic began a voluntary transition from CCA to other wood preservatives for certain residential uses such as play structures, picnic tables, decks, fencing, and boardwalks. This phase out was completed on December 31, 2003; however, wood treated prior to this date could still be used and existing structures made with CCA-treated wood would not be affected. CCA-treated wood products continue to be used in industrial applications. It is not known whether, or to what extent, CCA treated wood products may contribute to exposure of people to arsenic (ATSDR, 2005a).

In the past, inorganic arsenic compounds were predominantly used as pesticides, primarily on cotton fields and in orchards. Inorganic arsenic compounds can no longer be used in agriculture. However, organic arsenic compounds, namely cacodylic acid, disodium methylarsenate (DSMA), and monosodium methylarsenate (MSMA), are still used as pesticides (ATSDR, 2005a), principally on cotton. Some organic arsenic compounds are used as additives in animal feed. Small quantities of arsenic metal are added to other metals to form metal mixtures or alloys with improved properties. The greatest use of arsenic in alloys is in lead-acid batteries for automobiles. Another important use of arsenic compounds is in semiconductors and light- emitting diodes (ATSDR, 2005a).

Because arsenic occurs naturally in soil and minerals it may enter the air, water, and land from wind-blown dust and may get into water from runoff and leaching. Arsenic also enters the environment during the mining and smelting of ores and is released into the atmosphere from coal-fired power plants and incinerators. Arsenic released from power plants and other combustion processes is usually attached to very small particles. Arsenic contained in wind- borne soil is generally found in larger particles. These particles settle to the ground or are washed out of the air by rain. Many common arsenic compounds can dissolve in water. Thus, arsenic can get into lakes, rivers, or underground water by dissolving in rain or snow or through the discharge of industrial wastes. Some of the arsenic will stick to particles in the water or sediment on the bottom of lakes or rivers, and some will be carried along by the water. Ultimately, most arsenic ends up in the soil or sediment. Although some fish and shellfish take in arsenic, which may build up in tissues, most of this arsenic is in an organic form called arsenobetaine (commonly called "fish arsenic") that is much less harmful (ATSDR, 2005a).

Humans are exposed to some arsenic by eating food, drinking water, or breathing air. Inorganic arsenic has been recognized as a human poison since ancient times, and large oral doses (above 60,000 ppb in food or water) can result in death. Swallowing of lower levels of inorganic arsenic (ranging from about 300 to 30,000 ppb in food or water), may cause irritation of the stomach and intestines. Perhaps the single-most characteristic effect of long-term oral exposure to inorganic arsenic is a pattern of skin changes. These include a darkening of the skin and the appearance of small "corns" or "warts" on the palms, soles, and torso, and are often associated with changes in the blood vessels of the skin. Swallowing arsenic has also been reported to increase the risk of cancer in the liver, bladder, kidneys, prostate, and lungs. The Department of Health and Human Services has determined that inorganic arsenic is known to be a human carcinogen. The International Agency for Research on Cancer (IARC) has determined that inorganic arsenic is carcinogenic to humans. The USEPA also has classified inorganic arsenic as a known human carcinogen (ATSDR, 2005a).

Bioconcentration of arsenic occurs in aquatic organisms, primarily in algae and lower invertebrates. Both bottom-feeding and predatory fish can accumulate contaminants found in water. Bottom-feeders are readily exposed to the greater quantities of metals that accumulate in sediments. Predators may bioaccumulate metals from the surrounding water or from feeding on other fish, including bottom feeders, which can result in the biomagnification of the metals in their tissues. An extensive study of the factors affecting bioaccumulation of arsenic in two streams in western Maryland in 1997-1998 found no evidence of biomagnification since arsenic concentrations in organisms tend to decrease with increasing trophic level (ATSDR, 2005a). Arsenic is mainly accumulated in the exoskeleton of invertebrates and in the livers of fish. No differences were found in the arsenic levels in different species of fish, which included herbivorous, insectivorous, and carnivorous species. The major bioaccumulation transfer is between water and algae, at the base of the food chain and this has a strong impact on the concentration in fish. National Contaminant Biomonitoring data produced by the Fish and Wildlife Service were used to test whether differences exist between bottom-feeders and predators in tissue levels of metals and other contaminants (ATSDR, 2005a). No differences were found for arsenic. Bioconcentration factors measured in freshwater invertebrates and fish for several arsenic compounds ranged from 0 to 17, but a BCF of 350 was observed in marine oysters. In a study conducted at the Times Beach Confined Disposal Facility in Buffalo, New York, arsenic concentrations in tissue from zebra mussels exposed for 34 days were significantly higher than water column concentrations (ATSDR, 2005a). Barnacles growing on CCA-treated wood docks accumulated arsenic. The highest concentrations of arsenic were found on the most recently treated wood. Biomagnification in aquatic food chains does not appear to be significant, although some fish and invertebrates contain high levels of arsenic compounds (ATSDR, 2005a).

IV.3.5 Chromium (Cr) Chromium is a naturally occurring element found in rocks, animals, plants, soil, and in volcanic dust and gases. The metal is a lustrous steel-gray color which takes a high polish. It is hard and resistant to corrosion. Chromium compounds are toxic. Chromium is present in the environment in several different forms. The most common forms are chromium (0), trivalent (or chromium (III)), and hexavalent (or chromium (VI)). Chromium (III) occurs naturally in the environment and is an essential nutrient required by the human body to promote the action of insulin in body tissues so that sugar, protein, and fat can be used by the body. Chromium(VI) and chromium(0)

are generally produced by industrial processes.. The naturally occurring mineral chromite in the chromium (III) form is used as brick lining for high-temperature industrial furnaces, for making metals and alloys, and chemical compounds. Chromium compounds, mostly in chromium (III) or chromium(VI) forms, produced by the chemical industry are used for chrome plating, the manufacture of dyes and pigments, leather tanning, and wood preserving (ATSDR, 2000c).

The health effects resulting from exposure to chromium (III) and chromium (VI) are fairly well described in the literature. In general, chromium (VI) is more toxic than chromium (III). Breathing in high levels (greater than 2 µg/m3) of chromium (VI) can cause irritation to the nose. Lung cancer may occur long after exposure to chromium has ended. Chromium (VI) is believed to be primarily responsible for the increased lung cancer rates observed in workers who were exposed to high levels of chromium in workroom air. Swallowing small amounts of chromium (VI) has few adverse health effects; however, accidental or intentional swallowing of larger amounts has caused stomach upsets and ulcers, convulsions, kidney and liver damage, and even death. The levels of chromium (VI) that caused these effects were far greater than those that one might be exposed to in food or water. Although chromium (III) in small amounts is a nutrient needed by the body, swallowing large amounts of chromium (III) may cause health problems..

Because some chromium (VI) compounds have been associated with lung cancer in workers and caused cancer in animals, the Department of Health and Human Services has determined that certain chromium (VI) compounds (calcium chromate, chromium trioxide, lead chromate, strontium chromate, and zinc chromate) are known human carcinogens. The International Agency for Research on Cancer (IARC) has determined that chromium (VI) is carcinogenic to humans, based on sufficient evidence in humans for the carcinogenicity of chromium (VI) compounds as found in chromate production, chromate pigment production, and chromium plating industries. IARC's determination is also based on sufficient evidence in experimental animals for the carcinogenicity of calcium chromate, zinc chromate, strontium chromate, and lead chromate; and limited evidence in experimental animals for the carcinogenicity of chromium trioxide (chromic acid) and sodium dichromate. IARC has also determined that chromium (0) and chromium (III) compounds are not classifiable as to their carcinogenicity to humans. The EPA has determined that chromium (VI) in air is a human carcinogen. The EPA has also determined that there is insufficient information to determine whether chromium (VI) in water or food and chromium (III) are human carcinogens (ATSDR, 2000c).

Most of the chromium in lakes and rivers will ultimately be deposited in the sediments. Chromium in the aquatic phase occurs in the soluble state or as suspended solids adsorbed onto clay, organics, or iron oxides. Most of the soluble chromium is present as chromium (VI) or as soluble chromium (III) complexes and generally accounts for a small percentage of the total. Soluble chromium (VI) may persist in some bodies of water for a long time, but will eventually be reduced to chromium (III) by organic matters or other reducing agents in water. The residence times of chromium (total) in lake water range from 4.6 to 18 years (ATSDR, 2000c). Chromium (III) in soil is mostly present as insoluble carbonate and oxide of chromium (III); therefore, it will not be mobile in soil. The solubility of chromium (III) in soil and its mobility may increase due to the formation of soluble complexes with organic matters in soil. A lower soil pH may facilitate complexation. Chromium has a low mobility for translocation from roots to the above ground parts of plants (ATSDR, 2000c).

Most of the chromium released into water will ultimately be deposited in the sediment. A very small percentage of chromium can be present in water in both soluble and insoluble forms. Soluble chromium generally accounts for a very small percentage of the total chromium. Less than 0.002% of total chromium in water and sediment in the Amazon and Yukon Rivers was present in a soluble form (ATSDR, 2000c). In the aquatic phase, chromium (III) occurs mostly as suspended solids adsorbed onto clayish materials, organics, or iron oxide (Fe2O3) present in water. Approximately 10.5–12.6% of chromium in the aquatic phase of the Amazon and Yukon rivers was in solution, the rest being present in the suspended solid phase. The ratio of suspended to dissolved solid in an organic-rich river in Brazil was 2.1 (ATSDR, 2000c). Chromium does enter the food chain and the BCF for chromium (VI) in rainbow trout is 1 (ATSDR, 2000c).

IV.3.6 Copper (Cu) Copper is a reddish metal that occurs naturally in rock, soil, water, sediment, and, at low levels, air. The reddish color of this element is most commonly seen in the U.S. penny, electrical wiring, and some water pipes. It is also found in many mixtures of metals, called alloys, such as brass and bronze. Copper is primarily used as the metal or alloy in the manufacture of wire, sheet metal, pipe, and other metal products. Copper compounds are most commonly used in agriculture to treat plant diseases, like mildew, or for water treatment and as preservatives for wood, leather, and fabrics (ATSDR, 2004).

Copper is an essential nutrient that is incorporated into a number of metalloenzymes involved in hemoglobin formation, drug/xenobiotic metabolism, carbohydrate metabolism, catecholamine biosynthesis, and the antioxidant defense mechanism. Symptoms associated with copper deficiency in humans include anemia, leukopenia, and osteoporosis; copper deficiency is rarely observed in the U.S. general population (ATSDR, 2004).

Copper is widespread in the environment since about 640,000 tons of copper were released into the environment by industries in 2000. Elemental copper does not break down in the environment. Copper can be found in plants and animals, and at high concentrations in filter feeders such as mussels and oysters (ATSDR, 2004). When copper and copper compounds are released into water, the copper that dissolves can be carried in surface waters either in the form of copper compounds or as free copper or, more likely, copper bound to particles suspended in the water. Even though copper binds strongly to suspended particles and sediments, there is evidence to suggest that some water-soluble copper compounds do enter groundwater. Copper that enters water eventually collects in the sediments of rivers, lakes, and estuaries. Copper binds primarily to organic matter in estuarine sediment, unless the sediment is low in organic matter content. In the water column and in sediments, copper adsorbs to organic matter, hydrous iron and manganese oxides, and clay. In the water column, a significant fraction of the copper is adsorbed within the first hour of introduction, and in most cases, equilibrium is obtained within 24 hours. In fact, most of the copper in treated sewage plant effluent and surface runoff is already in the form of complexes (ATSDR, 2004).

The bioconcentration of copper in fish obtained in field studies indicate a low potential for bioconcentration. The BCF is higher in mollusks, such as oysters, and squid where it may reach 3.0x104 and 2.1x107, respectively (ATSDR, 2004).

IV.3.7 Nickel (Ni) Pure nickel is a hard, silvery-white metal that has properties that make it very desirable for combining with other metals to form alloys. Some of the metals that nickel can be alloyed with are iron, copper, chromium, and zinc. These alloys are used in making metal coins and jewelry and in industry for making items such as valves and heat exchangers. Most nickel is used to make stainless steel. There are also compounds consisting of nickel combined with many other elements, including chlorine, sulfur, and oxygen. Many of these nickel compounds are water soluble and have a characteristic green color. Nickel compounds are used for nickel plating, to color ceramics, to make some batteries, and as substances known as catalysts that increase the rate of chemical reactions. Nickel combined with other elements occurs naturally in the earth's crust. It is found in all soil, and is also emitted from volcanoes. In the environment, it is primarily found combined with oxygen or sulfur as oxides or sulfides. Nickel is released into the atmosphere during nickel mining and by industries that make or use nickel, nickel alloys, or nickel compounds. These industries also might discharge nickel in wastewater. Nickel is also released into the atmosphere by oil-burning power plants, coal-burning power plants, and trash incinerators. A lot of nickel released into the environment ends up in soil or sediment where it strongly attaches to particles containing iron or manganese. Under acidic conditions, nickel is more mobile in soil and might seep into groundwater. Nickel does not appear to concentrate in fish. Studies show that some plants can take up and accumulate nickel. However, it has been shown that nickel does not accumulate in small animals living on land that has been treated with nickel-containing sludge. The most common harmful health effect of nickel in humans is an allergic reaction. Approximately 10 - 20% of the population is sensitive to nickel. People who are not sensitive to nickel must eat very large amounts of nickel to suffer harmful health effects (ATSDR, 2005d).

Typical concentrations of nickel reported in soil range from 4 to 80 ppm. Nickel may be transported into streams and waterways from the natural weathering of soil as well as from anthropogenic discharges and runoff. This nickel accumulates in sediment. Nickel levels in surface water are generally low. In some studies, nickel could not be detected in a large fraction of analyzed samples (ATSDR, 2005d). Median nickel concentrations in seawater are typically 0.1 - 0.5 μg/L. The speciation and physicochemical state of nickel is important in considering its behavior in the environment and availability to biota. For example, the nickel incorporated in some mineral lattices may be inert and have no ecological significance. Most analytical methods for nickel do not distinguish the form of nickel; the total amount of nickel is reported, but the nature of the nickel compounds and whether they are adsorbed to other material is not known. This information, which is critical in determining nickel's liability and availability, is site specific. Therefore, it is impossible to predict nickel's environmental behavior on a general basis. In soil, the most important sinks for nickel, other than soil minerals, are amorphous oxides of iron and manganese. The mobility of nickel in soil is site specific depending mainly on soil type and pH. The mobility of nickel in soil is increased at low pH. At one well-studied site, the sulfate concentration and the surface area of soil iron oxides were also key factors affecting nickel adsorption (ATSDR, 2005d).

Nickel in sediment may be strongly bound or present in a removable form. The nickel concentration in 450 uncontaminated estuarine and coastal marine sites in the southeastern United States covaried significantly with the aluminum concentration, suggesting that natural

aluminosilicates are the dominant natural metal-bearing phase in some aquatic systems. Nickel removed by coprecipitation can be remobilized by microbial action under anaerobic conditions. Remobilization results from enzymatic reductive dissolution of iron with subsequent release of coprecipitated metals (ATSDR, 2005d).

IV.3.8 Zinc (Zn) Zinc is one of the most common elements in the Earth's crust. Zinc is found in the air, soil, and water and is present in all foods. In its pure elemental (or metallic) form, zinc is a bluish-white, shiny metal. Metallic zinc has many uses in industry. Perhaps the most common use for zinc is to coat steel and iron as well as other metals to prevent rust and corrosion; this process is called galvanization (ATSDR, 2005e).

Zinc enters the air, water, and soil as a result of both natural processes and human activities. Most zinc enters the environment as the result of mining, purifying of zinc, lead, and cadmium ores, steel production, coal burning, and burning of wastes. Waste streams from zinc and other metal manufacturing and zinc chemical industries, domestic wastewater, and run-off from soil containing zinc can discharge zinc into waterways. The level of zinc in soil increases mainly from disposal of zinc wastes from metal manufacturing industries and coal ash from electric utilities. Sludge and fertilizer also contribute to increased levels of zinc in the soil. In air, zinc is present mostly as fine dust particles. This dust eventually settles over land and water. Rain and snow aid in removing zinc from air. Most of the zinc in lakes or rivers settles on the bottom. However, a small amount may remain either dissolved in water or as fine suspended particles. The level of dissolved zinc in water may increase as the acidity of water increases. Fish can collect zinc in their bodies from the water they swim in and from the food they eat. Most of the zinc in soil is bound to the soil and does not dissolve in water. However, depending on the type of soil, some zinc may reach groundwater, and contamination of groundwater has occurred from hazardous waste sites. Zinc may be taken up by animals eating soil or drinking water containing zinc. Zinc is also a trace mineral nutrient and as such, small amounts of zinc are needed in all animals (ATSDR, 2005e).

Zinc is an essential element needed by humans in small amounts. Taking too much zinc into the body through food, water, or dietary supplements can affect health negatively. If large doses of zinc (10 - 15 times higher than the RDA) are taken by mouth even for a short time, stomach cramps, nausea, and vomiting may occur. Ingesting high levels of zinc for several months may cause anemia, damage the pancreas, and decrease levels of high-density lipoprotein (HDL) cholesterol. The USEPA has determined that because of lack of information, zinc is not classifiable as to its human carcinogenicity (ATSDR, 2005e).

Zinc occurs in the environment mainly in the +2 oxidation state and sorption is the dominant reaction, resulting in the enrichment of zinc in suspended and bed sediments. Zinc in aerobic waters is partitioned into sediments through sorption onto hydrous iron and manganese oxides, clay minerals, and organic material. The efficiency of these materials in removing zinc from solution varies according to their concentrations, pH, redox potential (Eh), salinity, nature and concentrations of complexing ligands, cation exchange capacity, and the concentration of zinc. Precipitation of soluble zinc compounds appears to be significant only under reducing conditions in highly polluted water. Generally, at lower pH values, zinc remains as a free ion. The free ion

(Zn+2) tends to be adsorbed and transported by suspended solids in unpolluted waters. In polluted waters in which the concentration of zinc is high, removal of zinc by precipitation of the hydroxide form is possible, particularly when the pH is >8. Estuaries like Bayou Grande with pH’s ~8 could favor precipitation. In anaerobic environments and in the presence of sulfide ions, precipitation of zinc sulfide limits the mobility of zinc. Although biota appear to be a minor reservoir of zinc relative to soils and sediments, microbial decomposition of biota in water can produce ligands, such as humic acids, that can affect the mobility of zinc in the aquatic environment through zinc precipitation and adsorption. Zinc concentrations in the air are relatively low, except near industrial sources such as smelters. No estimate for the atmospheric lifetime of zinc is available at this time, but the fact that zinc is transported long distances in air indicates that its lifetime in air is at least on the order of days (ATSDR, 2005e).

Zinc can accumulate in freshwater animals at 51 - 1,130 times the concentration present in the water. Microcosm studies indicate, in general, that zinc does not biomagnify through food chains (ATSDR, 2005e). Steady state zinc BCFs for 12 aquatic species range from 4 to 24,000. Crustaceans and fish can accumulate zinc from both water and food. A BCF of 1,000 was reported for both aquatic plants and fish, and a value of 10,000 was reported for aquatic invertebrates. Other investigators have also indicated that organisms associated with sediments have higher zinc concentrations than organisms living in the water column (ATSDR, 2005e).

V OBJECTIVES

The general goal of this PERCH project was to evaluate the presence, magnitude and potential origin of pollutants in Bayou Grande. Specifically, the project was designed to address the following objectives: • Review all accessible information related to pollution of Bayou Grande and its drainage basin to identify data gaps and areas of concern. This objective includes: o Determining historical impacts o Evaluating earlier studies o Assessing the most recent studies at NAS Pensacola • Characterize selected SOCs of water and sediments in Bayou Grande, which includes: o Determining the presence and magnitude of the SOCs o Assessing the source and mode of transport of he SOCs o Evaluating if SOCs are currently entering Bayou Grande o Appraising the rate of degradation of existing SOCs o Establishing relationships between pollution and sediment characteristics • Assess and characterize dioxins/furans and dioxin-like PCBs in sediments in Bayou Grande in support of a related PERCH Seafood Study.

VI METHODS

Accessible information concerning the environmental conditions of Bayou Grande was compiled through an exhaustive literature search. For this effort we drew in part upon another component of the PERCH Project, the PERCH Bibliography (http://fusionmx.lib.uwf.edu/perch/index.cfm) which is a fully searchable database of bibliographical materials pertaining to the environment of Northwest Florida. A GIS database of spatially referenced data collected during the literature search was constructed by manually entering and digitally importing the data and by converting them to common spatial parameters. The purpose of the literature search was to assess what was known about the environmental quality of Bayou Grande and if it was or could be impacted by superfund sites and other potential sources of pollution. This information allowed evaluation of how the present project could further the existing knowledge. To help identify optimal locations for the sampling sites the bathymetry of the Bayou was surveyed with an echosounder and differential GPS (DGPS). Optimal sampling locations were identified by project personnel based on the bathymetry of the Bayou, the general location within the Bayou, and specific objectives of the sampling. These locations were marked on an overlay on the bathymetric map in the GIS. In the field, the GIS was used in combination with a WAAS enabled hand held GPS receiver (Garmin GPS V) to navigate to the sampling locations.

The sampling was conducted from fall of 2006 to mid-summer of 2007. A total of 78 sediment grab samples, 8 water grabs, and 10 vibracores were collected (Figure 19, 20 and 21). Sediment grab samples were collected with a ponar grab from a small boat. Water grabs were collected with a Van Dorn sampler. For the vibracores three-inch decontaminated aluminum thin-walled irrigation pipe was clamped to a vibracore powered by a portable generator. The vibracore sediment was retained by a plastic core catcher at the bottom and a vacuum plug sealed the top upon retrieval of the coring pipe. Each vibracore was driven into the sediment until refusal. In the lab the cores were split lengthwise, described, and sampled at 1 meter intervals (0 meters for level A, one meter for level B, two meters for level C, and 3 meters for level D). For the sediment grab samples five local grab samples were joined at each sampling site and mixed thoroughly prior to further processing. The composited samples were placed into dedicated sampling containers and sent to the analytical laboratory the day of sampling. Sampling equipment was cleaned with soapy water, rinsed with reagent grade solvents, and two rinses of HPLC grade water. The decontaminated equipment was tested with rinsate blanks, and field splits for quality control were taken. The first sampling series, the GBc series of sediment grabs, was taken in embayments of the Bayou and along the shoreline near possible stormwater outfalls. GBc samples were analyzed for trace metals and PAH. Sample series GF was taken elsewhere in the Bayou and was analyzed for PAH, trace metals, pesticides, PCBs, total petroleum hydrocarbons, and dioxins/furans.

It should be borne in mind that Ponar dredge surface grabs and the Level A (“surface level”) of the vibracores are not equivalent because the sediment depths and the homogeneity of sediments sampled differ. Ponar grab samplers like the one used in the current study do not collect deeper than 13 cm, resulting in data that reflect concentrations in the surface sediments. Vibracore samples are collected from sediments in a 3-inch diameter pipe. Sampling starts at the surface and continues down from the surface to depth of about 20 cm, or until sufficient sediment has

been collected for analyses. Thus, a level A vibracore sample (0 m) differs from surface grab samples relative to the depth of the sampled sediments. The surface area sampled is for a vibracore is about 46 cm2 versus about 256 cm2 for a petite ponar dredge. The dredge sample additionally consists of three to five composited samples. The smaller surface area of vibracores gives a more discrete result that is more likely to vary from the mean sediment analyte concentration due to the heterogeneous distribution of nonsoluble materials in sediments.

Figure 19. Location of sediment grab samples.

Figure 20. Location of water grab samples.

Figure 21. Location of vibracore sites in Bayou Grande.

Analytical methods followed standard procedures. Total petroleum was analyzed by the FDEP FL-PRO method. The FL-PRO analysis is designed to measure concentrations of total petroleum hydrocarbons (TPH) in water and soil/sediment in the alkane analytical range of C8-C40. The method is based on a solvent extraction and gas chromatography procedure (using a Flame Ionization Detector). Silica cleanup is a mandatory part of the procedure, designed to remove potential interferences from animal and vegetable oil and grease and biogenic terpenes. Other organic compounds, including chlorinated hydrocarbons, phenols and phthalate esters are detected and the total concentration values of TPH for the FL-PRO may include these compounds in the results. USEPA SW-846 methods were used for the following: PCBs by 1668A, dioxins/furans by 1613B, and other semivolatiles by Method 8270C. Specific PAHs (naphthalene, 2-methylnaphthalene, 1-methylnaphthalene, acenaphthylene, acenaphthene, fluorene, phenanthrene, anthracene, fluoranthene, pyrene, chrysene, benz(a)anthracene, benzo(b)fluoranthene, benzo(k)fluoranthene, benzo(a)pyrene, indeno(1,2,3-cd)pyrene, dibenz(a,h)anthracene, and benzo(g,h,i)perylene) were analyzed by USEPA SW-846 method 8270 C, with Simultaneous Ion Monitoring (SIM). This method was used to achieve detection (MDL) and reporting limits (RL) that were lower than the Florida marine sediment quality assessment guidelines (SQAGs) (MacDonald, 1994a, b). SIM is the most sensitive gas chromatography/ mass spectrometry method that is generally available for PAH detection. The target analytes are extracted by EPA method 3550 using dichloroethane (methylene chloride), separated by gas chromatography, then identified and quantitated by mass spectrometry. SIM is a method in which the detector lingers at a few selected masses for much longer than when using the typical "full scan mode", thus increasing the sensitivity of the detector to those masses and lowering both the method detection limit (MDL) and reporting limit (RL) for the analytes. Organochlorinated pesticides were analyzed by EPA Method 8081A which is a gas chromatography method that is similar to EPA Method 8082 that is used to detect PCBs. It employs fused-silica, open tubular capillary columns with electron capture detection. The 8081 analyses were run for the following pesticides: alpha-BHC, gamma-BHC (Lindane), beta-BHC, delta-BHC, heptachlor, aldrin, heptachlor epoxide, gamma-chlordane, alpha-chlordane, 4,4'- DDE, endosulfan I, dieldrin, endrin, 4,4'-DDD, endosulfan II, 4,4'-DDT, endrin aldehyde, methoxychlor, endosulfan sulfate, endrin ketone, toxaphene, tetrachloro-m-xylene, decachlorobiphenyl. The EPA Method 3550 was also used for the extraction.

Mercury (Hg) was determined by Method 7471A for sediments and Method 7470A for aqueous samples by cold vapor atomic absorption. For all other trace metal determinations the samples were prepared according to SW-846 Method 6010, Acid Digestion of Sediments, Sludges, and Soils. Per the method, aluminum (Al), calcium (Ca), iron (Fe), magnesium (Mg), nickel (Ni), selenium (Se), tin (Sn), cadmium (Cd), copper (Cu), zinc (Zn), arsenic (As), chromium (Cr), and lead (Pb) were prepared for graphite furnace atomic absorption spectrometry (GFAAS). The other metals were prepared for flame atomic absorption spectrometry (FLAAS). The digestates were analyzed according to Standard Method 3111 for FLAAS or USEPA Method 200.9 for GFAAS. Samples for particle size analysis were manually mixed and homogenized in the lab while being air dried. After air drying, samples were crushed with mortar and pestle to break up aggregates. Analyses were then performed by dry, Ro-tap, sieving for the sand fractions (2 mm - 0.063 mm) and by the pipette method for clays (procedure 3A1 of Burt (2004)). We preferred to use the pipette method over the often employed hydrometer method because the pipette method is generally considered to be more accurate.

The metal, volatile, total organic carbon, and semivolatile analysis were performed by Columbia Analytical Systems of Jacksonville, FL, and their high resolution mass spectrometry laboratory in Houston, TX, performed the analyses for PCB and dioxin/furan congeners. Particle size analyses were performed at the Sediments Lab, Department of Environmental Studies, University of West Florida.

To calculate a TEQ for the dioxins/furans and dioxin-like PCBs the TEF of each congener present in a mixture was multiplied by the respective mass concentration and the products were summed to represent the 2,3,7,8-TCDD TEQ of the mixture, as determined by the following equation (USEPA, 2003b):

TEQ ≅ Σi−n(Congener i × TEF i)+(Congener j × TEF j) + ....+ (Congener n × TEF n)

The TEF values used were those for humans/mammals established in 2005 by the WHO (Van den Berg et al., 2006) (Table 13).

To assess the origin of dioxins/furans, dioxin-like PCBs and PCBs we statistically examined similarities in their profiles with principal component analysis (PCA) and cluster analysis. We applied a varimax rotated PCA to facilitate interpretation of the resulting components.

Table 13. TEF values for dioxins/furans and dioxin-like PCBs [ng/kg toxic equivalents of 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD)] (Van den Berg et al., 2006).

Compound WHO 1998 TEF WHO 2005 TEF Chlorinated dibenzo-p-dioxins 2,3,7,8-TCDD 1 1 1,2,3,7,8-PeCDD 1 1 1,2,3,4,7,8-HxCDD 0.1 0.1 1,2,3,6,7,8-HxCDD 0.1 0.1 1,2,3,7,8,9-HxCDD 0.1 0.1 1,2,3,4,6,7,8-HpCDD 0.01 0.01 OCDD 0.0001 0.0003 Chlorinated dibenzofurans 2,3,7,8-TCDF 0.1 0.1 1,2,3,7,8-PeCDF 0.05 0.03 2,3,4,7,8-PeCDF 0.5 0.3 1,2,3,4,7,8-HxCDF 0.1 0.1 1,2,3,6,7,8-HxCDF 0.1 0.1 1,2,3,7,8,9-HxCDF 0.1 0.1 2,3,4,6,7,8-HxCDF 0.1 0.1 1,2,3,4,6,7,8-HpCDF 0.01 0.01 1,2,3,4,7,8,9-HpCDF 0.01 0.01 OCDF 0.0001 0.0003 Non-ortho–substituted PCBs 3,3',4,4'-tetraCB (PCB 77) 0.0001 0.0001 3,4,4',5-tetraCB (PCB 81) 0.0001 0.0003 3,3',4,4',5-pentaCB (PCB 126) 0.1 0.1 3,3',4,4',5,5'-hexaCB (PCB 169) 0.01 0.03 Mono-ortho–substituted PCBs 2,3,3',4,4'-pentaCB (PCB 105) 0.0001 0.00003 2,3,4,4',5-pentaCB (PCB 114) 0.0005 0.00003 2,3',4,4',5-pentaCB (PCB 118) 0.0001 0.00003 2',3,4,4',5-pentaCB (PCB 123) 0.0001 0.00003 2,3,3',4,4',5-hexaCB (PCB 156) 0.0005 0.00003 2,3,3',4,4',5'-hexaCB (PCB 157) 0.0005 0.00003 2,3',4,4',5,5'-hexaCB (PCB 167) 0.00001 0.00003 2,3,3',4,4',5,5'-heptaCB (PCB 189) 0.0001 0.00003

VII RESULTS AND DISCUSSION

VII.1 Semivolatile Organic Compounds

VII.1.1 Total Petroleum Hydrocarbons Bayou Grande has NAS on its southern bank. NAS is a major center for US Navy aviation activities that consumes large quantities of petroleum products that include aviation fuels, petroleum fuels, other products for vehicles and marine craft, and other diverse demands for petroleum products. NAS and also the urban activities on the northern shore of the Bayou are likely sources of spills related to fuel consumption. A likely mode of transport to the bayou of spilled petroleum product would involve stormwater and direct releases into the waters from marine activities. Contaminated aquifers are another possible source for TPH.

There were four detections for TPH out of 23 analyzed sediment samples (Table 14). Three of these detections occurred in sediments located in embayments that are adjacent to the more developed areas of the watershed (Figure 22). All of the detections were below the analytical reporting limit for the analyses. The highest detection (sample GF-17, 150 mg/kg) was in Davenport Bayou that is actually not part of Bayou Grande, but is directly adjacent to it. The second highest detected level of petroleum hydrocarbons was from sample GF-2 (69 mg/kg) in the Yacht Basin of NAS, where sailing boats are moored. Detections also occurred for sample GF-19 taken from Navy Point Bayou (also adjacent to Warrington) and GF-20, which is located in the western end of the bayou that is proximal to the most undeveloped regions of Bayou Grande. In Bayous Texar and Chico, a short distance to the east of Bayou Grande, the frequency of detections and the concentrations of TPH were higher with maximum values up to 1700 mg/kg in Bayou Texar (Mohrherr et al., 2005; 2006). The method detection limit (MDL) for Bayou Grande sediment for these analyses was approximately 8-130 mg/kg. It could be argued that a lower MDL would have resulted in more detections, however the highest detected concentration of 150 mg/kg is several times less than the higher concentrations detected in Bayous Chico and Texar. These data indicate that the TPH concentrations in sediments in Bayou Grande are generally low.

The presence of TPH in the water column is, however, a source for concern because oil in aquatic systems is harmful to shellfish, finfish, marine mammals and waterfowl that live near the spill. Oil spills are unsightly to the general public and are expensive to clean up. In addition, damage to fisheries places a hardship on those who make their living by fishing. The concentrations of detected TPH were relatively low in water (highest was 0.19 mg/l) and no oil slick was sighted. Detection of TPH in water occurred in three out of eight samples. Aqueous sample GW-6 (located at the mouth of Bayou Grande), GW-7 (located in the Yacht Basin), and GW 8 (located in Redoubt Bayou) had trace amounts of petroleum (Table 15, Figure 20). It is not clear where the TPH detected in the three positive aqueous samples may have come from. Samples GW-7 and GW-8 in Redoubt Bayou are adjacent to likely sources but GW-6 is actually at the interface between Bayou Grande and Pensacola Bay. The non-detect samples were taken on February 6, 2007 and the three positive samples were taken on February 14, 2007. This suggests that a release of petroleum may have occurred between these dates. On February 13th,

just prior to the collection of February 14th, a 3.00 cm rain event occurred. This was the first rain after a 1.4 cm rain two weeks earlier. It is possible that the rainfall created runoff that transported a land based oil release to the bayou. Since it was found in three separate sites, it is not possible to predict a precise location for such a release.

Table 14. Total petroleum hydrocarbons [mg/kg] in surface sediments. Sample ID TPH GF-1 <331 GF-2 69 i2 GF-3 <57 GF-4 <73 GF-5 <110 GF-6 <91 GF-7 <110 GF-8 <96 GF-9 <92 GF-10 <7.9 GF-11 <8 GF-12 <8 GF-13 <37 GF-14 <140 GF-15 <91 GF-16 <26 GF-17 150 i GF-18 <31 GF-19 14 i GF-20 54 i GF -21 <110 GF -22 <120 GF -23 <130 1: < indicates a non-detect showing that the result is below the listed MDL (method detection limit). 2: i is a data qualifier and indicates that the reported value is between the laboratory method detection limit and the laboratory practical quantitation limit.

Figure 22. Total petroleum hydrocarbons in sediments.

Table 15. Total petroleum hydrocarbons [mg/l] in water. Sample TPH GW-1 <0.0621 GW-2 <0.062 GW-3 <0.062 GW-4 <0.062 GW-5 <0.062 GW-6 0.19 i2 GW-7 0.15 i GW-8 0.083 i 1: < indicates a non-detect showing that the result is below the listed MDL. 2: i is a data qualifier and indicates that the reported value is between the laboratory method detection limit and the laboratory practical quantitation limit.

Currently there are no applicable sediment quality guidelines for this general range of hydrocarbons. A discussion under the auspices of the USACE (US Army Corps of Engineers) Seattle District commented upon this issue. Thornburg (2004) found that screening levels for bulk petroleum hydrocarbons in sediment have not been developed due to the widely varying

mix of compounds that can be present in total petroleum and a perception that toxicity could be adequately accounted for by considering the toxicity of PAHs that are a constituent of total petroleum. However, there are situations where bulk petroleum hydrocarbons are present in sediment at elevated levels, and individual listed constituents either are absent or are present at levels that would not indicate toxicity. Total petroleum hydrocarbon analyses are routinely performed, but it is difficult to relate directly any specific concentration to toxicity. This is due to there having been limited study of the relationship between toxicity threats to sediments by petroleum contamination and the general diversity and variability of TPH constituents. TPH toxicity or other related detrimental effects upon the bayou are not clear and this is reflected in what appears to be an absence of federal or FDEP recognized marine sediment quality assessment guidelines (SQAGs) for total petroleum. More recent theories for assessing the toxicity of petroleum and its constituents to benthic organisms have focused on a narcosis-based approach (Abernathy et al., 1988; Franks and Lieb, 1978).

VII.1.2 Polycyclic Aromatic Hydrocarbons (PAHs)

VII.1.2.1 PAHs in surface sediments Presently there are only sediment guidelines (TEL and PEL) for 13 of the PAH species. These 13 species comprise only a small number of the PAHs that commonly occur in the environment. The approved USEPA analytical methods used in the present study are set up to detect 18 different PAH species. We assume that there are undetected PAH species, but have no direct proof that they are present. Aromatics (including PAHs) are considered to be the most acutely toxic component of petroleum products, and are also associated with chronic and carcinogenic effects. Sixteen of the common PAHs typically analyzed in standard laboratory scans have been listed by the Environmental Protection Agency among 126 priority pollutants under the Clean Water Act (Federal Register, 1997). Four of them (benzo(a)pyrene, benzo(b)fluoranthene, dibenzo(a,h)anthracene, and benzidine) are also listed among the 25 most hazardous substances thought to pose the most significant potential threat to human health at Superfund Sites (Van Mouwerik et al., 1998). There are also IARC (International Agency for Research on Cancer) listings for cancer causing PAHs (Table 3).

All 18 PAH species detectable via the standard 8270 C SIM analysis were detected in Bayou Grande: naphthalene, 2-methylnaphthalene, 1-methylnaphthalene, Acenaphthylene, Acenaphthene, Fluorene, phenanthrene, anthracene, fluoranthene, pyrene, chrysene, Benz(a)anthracene, benzo(b)fluoranthene, benzo(k)fluoranthene, benzo(a)pyrene, Indeno(1,2,3- cd)pyrene, dibenzo(a,h)anthracene, and benzo(g,h,i)perylene. A major concern of the project was to achieve analytical results that would give MDLs (method detection limits) and RLs (reporting limits) below the FDEP TEL. Table 16 shows the MDL (method detection limit) and RL for PAH analytes analyzed by 8270 C SIM for sample GBc-30, a representative sample. The MDLs for most analyses were lower than the TEL but more contaminated sediments had a higher MDL and RL than less contaminated sediments.

Table 16. Typical PAH minimal detection limit (MDL) and reporting limit (RL) and SQAGs (from sample GBc-30). 1 PAH Compound RL MDL2 TEL ug/kg PEL ug/kg LMW (Light Molecular Weight) PAHs Acenaphthene 8.7 3.5 6.71 88.9 Acenaphthylene 8.7 3.4 5.87 128 Anthracene 4.1 1.47 46.9 245 Fluorene 4.4 2 21.2 144 2-methylnaphthalene 4.4 0.2 20.2 201 Naphthalene 4.4 0.65 34.6 391 Phenanthrene 8.71 4.3 86.7 544 Sum LMW-PAHs na na 312 1,442 HMW (Heavy Molecular Weight) PAHs Benz(a)anthracene 4.4 0.64 74.8 693 4.4 Benzo(a)pyrene 1.5 88.8 763 4.4 Chrysene 0.62 108 846 4.4 Dibenzo(a,h)anthracene 0.65 6.22 135 4.4 Fluoranthene 0.76 113 1,494 4.4 Pyrene 0.67 153 1,398 Sum HMW-PAHs na na 655 6676 Sum LMW&HMW na na 1684 16,770 PAHs not assigned SQAG by FDEP Benzo(b)fluoranthene 4.1 1.6 NA NA Benzo(g,h,i)perylene 4.1 1.5 NA NA Benzo(k)fluoranthene 4.1 2.2 NA NA Indeno(1,2,3-cd)pyrene 4.1 1.5 NA NA 1-Methylnaphthalene 4.1 0.75 NA NA

The highest PAH concentration was detected in sample GBc-47 in the Yacht Basin with a total concentration of 101,730 ug/kg (Table 17, Figure 23) This sample exceeded FDEP PELs except for acenaphthene. Samples GBc-48 and GF-2 were also taken in the Yacht Basin. For GBc-48 PAH concentrations were below FDEP SQAGs (Table 17 and Figure 23). In sample GF-2 SQAG guidelines were exceeded (Table 18, Figure 23) but the concentrations of the PAHs were generally lower than in GBc-47. The overall PAH profile of GF-2 was strongly weighted towards LMW PAHs (56% of total PAHs were LMW, 2-methylnaphthalene alone was 19%). This latter PAH is associated with fuel spills, but can also have other origins. Surface grabs showed elevated levels for various SOCs in the Yacht Basin (see below). The lack of fine

sediments below the surface along with the likelihood that dredging took place in the basin (EnSafe, 2003; 2004) suggests that the underlying sediments may not be as polluted as the surface sediments. An attempt was made to obtain a vibracore in this embayment but without success because of the presence of compacted sand. The high levels of petroleum compounds that include PAHs, and other SOCs, can impact deeper sediments via contaminated groundwater transport. The presence of high 2-methylnaphthalene supports that possibility, but without vibracore data it is difficult to determine the source of the LMW PAHs in this embayment.

Navy Point Bayou above the Sunset Bridge showed the second highest concentration of PAHs (Figure 23, Table 17). Both duplicates of GBc-52 exceeded FDEP SQAG guidelines for most PAHs with the exception of four of the lighter molecular weight PAHs: naphthalene and 2- methyl naphthalene, acenaphthene, and acenaphthylene that were below the detection limits of that analytical run. The concentrations of the three remaining light molecular weight PAHs were sufficient to also make the total for LMW PAHs exceed the FDEP PEL. Other samples in that embayment show lower concentrations of PAHs (Figure 23, Figure 24) and none of these other samples exceeded the TEL for the Sum of LMW & HMW PAHs. Sample GF-15 taken in the proximity of a large culvert in Navy Point Bayou that was high for total PCB (see below) had a total PAH concentration of 1,897 ug/kg and exceeded the TEL for one LMW PAH and for all HMW PAHs (Figure 24, Table 18). Sample GBc-53 was very low in total PAH concentration but still exceeded the TEL for two HMW PAHs and one LMW PAH. Duplicate samples GBc- 51A and GBc-51B had concentrations that were below FDEP SQAGs. GF-19 in Navy Point Bayou also had very low concentrations and was below the SQAGs except for acenaphthene that exceeded the TEL. The only surface sample in Navy Point Bayou that had detections of a naphthalene was GF-23 (2-methylnaphthalene, 36 ug/kg, TEL = 20.2 ug/kg). This is likely due to its greater proximity to the main channel of the Bayou. Overall, PAHs in Navy Point Bayou surface sediments show no noticeable impact from fuel release, unlike in the Yacht Basin embayment. Davenport Bayou drains parts of Warrington that are next to the areas drained by Navy Point Bayou. In Davenport Bayou two out of three samples exceed the TEL concentration for the Sum of LMW & HMW PAHs. Sample GF-17 near the mouth of Davenport Bayou showed exceedances of SQAQs for naphthalenes and other LMW PAHs, and TPH was also detected at this site. Upstream in this embayment (GBc-49 and GBc-50) the lighter PAHs decrease in concentration while the proportion of heavier PAHs increases. The LMW PAHs near the mouth could originate from upstream in the embayment or could possibly have been carried in by tides and currents.

Other samples with PAH concentrations above SQAGs were in Woolsey Bayou and close to the shoreline between Redoubt Bayou and Woolsey Bayou. At the end of Woolsey Bayou GF-3 exceeded SQAGs (5,305 ug/kg total PAH), especially for the LMW PAHs (Figure 24, Table 18). Samples GBc-44 and 45, also taken at the end of Woolsey Bayou, exceeded the PEL for HMW PAHs and sample GBc-44 exceeded either the TEL or PEL for five LMW PAHs, but detections of naphthalenes show minimal concentrations.

Overall PAHs were highest in embayments that were closest to either NAS or dense urban areas. The main basin of Bayou Grande had lesser, but still significant, concentrations of PAHs. Only a few of these samples exceeded the SQAG for the Sum of LMW & HMW PAHs. Sample GF-1 in Pensacola Bay, adjacent to Bayou Grande, exceeded the SQAGs for LMW PAHs. In contrast,

GF-16 taken in Bayou Grande somewhat more than a kilometer west of GF-1 had concentrations below SQAGs. The presence of LMW PAHs in GF-1 suggests a relationship with nearby GF-17 in Davenport Bayou that has similar analyte concentrations. It is possible that the PAHs and other SOCs in these two samples could have origins from elsewhere other than Warrington such as the downtown Pensacola shoreline and the shipping lanes. The samples taken in the main bayou west of GF-16 (upstream) have higher total PAH concentrations than what was detected at GF-16 and exceeded some SQAGs for LMW PAHs (Figure 23). In the upper reach of the Bayou there was an overall decline for the detected PAH concentrations, but SQAGs were still exceeded for some PAHs (Tables 17, Table 18; Figure 23, Figure 24). The GF series of samples that were taken in the deeper portions of the Bayou frequently exceeded SQAGs for the naphthalene LMW PAHs that are present in petroleum fuels. Samples with exceedances for these occurred in 16 out of the 23 GF samples (Table 18).

Table 17. PAH concentrations [ug/kg] in sediments, GBc series. ID Naph1 2-Men 1-Men Acen Athen Fluor Phen Anth LMW Flrant TEL 34.62 20.2 5.9 6.7 21.2 86.7 46.9 544.0 113 PEL 391.03 201.0 128.0 88.9 144.0 544.0 245.0 1442.0 1494 1 <0.644 <0.8 <0.79 <0.84 <1.3 <0.77 <0.68 <0.86 0.0 <0.6 2 <0.64 <0.8 <0.79 2.4 i <1.3 <0.76 2.2 i <0.85 4.6 8 3 <0.64 <0.8 <0.79 <0.84 <1.3 <0.76 <0.67 <0.85 0.0 <0.6 4 <0.66 <0.83 <0.82 10.0 <1.4 <0.79 7 8.4 25.4 33 5 <0.61 <0.77 <0.75 <0.8 <1.3 <0.73 <0.65 <0.82 0.0 <0.57 6 <0.7 <0.88 <0.86 2.4 i <1.4 <0.83 3.8 i 1.7 i 7.9 28 7 <1.5 <1.9 <1.9 <2 <3 <1.8 10 i <2.1 10.0 37 8 <0.65 <0.82 <0.81 <0.86 <1.3 <0.78 <0.69 <0.87 0.0 2.9 i 9 <1.6 <2 <2 <2.1 <3.1 <1.9 4 i 2.5 i 6.5 19 10 <1.3 <1.6 <1.6 <1.7 <2.6 <1.6 3.7 i <1.7 3.7 10 11 <1.3 <1.6 <1.6 15.0 <2.5 <1.5 6.4 i 11 32.4 43 12 <0.7 <2.1 <1.6 <3.6 <3.7 <2.1 5.4 i <0.84 5.4 9.5 13 <0.85 <2.5 <1.9 <4.4 <4.5 <2.5 <5.5 <1.1 0.0 13 14 <1.6 <4.6 <3.4 <7.9 <8.2 <4.6 <10 <1.9 0.0 11 i 15 <1.5 <4.4 <3.3 <7.7 <8 <4.4 <9.7 <1.8 0.0 11 16 <1.3 <3.7 <2.7 <6.4 <6.6 <3.7 <8.1 <1.5 0.0 12 17 <0.72 <2.1 <1.6 <3.7 <3.8 <2.1 <4.7 <0.86 0.0 3.4 i 18 <0.67 <2 <1.5 <3.4 <3.6 <2 <1.2 <0.67 0.0 <0.85 19 <0.87 <2.6 <1.9 <4.5 <4.6 <2.6 <5.7 <1.1 0.0 6.3 20 7.8 i 11 5.6 i 6.7 i <1.3 <3.7 <6.3 <6.6 25.5 <3.7 21 4.5 i 4.9 3.9 i 3 i <0.72 <2.2 <3.7 <3.9 12.4 <2.2 22 10 9.8 6.8 4.8 i <0.83 <2.5 <4.3 <4.4 24.6 <2.5 23 2.9 i 2.9 i 2.3 i 1.8 i <0.7 <2.1 <3.6 <3.7 7.6 <2.1 24 <0.8 2.1 i 2.1 i <0.68 <0.7 <2.1 <3.6 <3.7 2.1 <2.1 25 <0.78 <0.69 <0.64 <0.66 <0.68 <2 <3.5 <3.6 0.0 <2 26 <0.64 <1.9 <1.4 <3.3 <3.4 <1.9 <4.2 <0.77 0.0 <0.74 27 <0.67 <2 <1.5 <3.4 <3.6 <2 <4.3 <0.8 0.0 5.1 28 <0.66 <2 <1.5 <3.4 <3.5 <2 <4.3 <0.78 0.0 <0.76 29 <0.66 <2 <1.5 <3.4 <3.5 <2 <4.3 <0.79 0.0 6.7 30 <0.65 <2 <1.5 <3.4 <3.5 <2 <4.3 <0.78 0.0 <0.76 31 <0.66 <2 <1.5 <3.4 <3.5 <2 <4.3 <0.79 0.0 <0.76 32 <2.4 <7 <5.1 <13 <13 <7 <16 <2.9 0.0 23 33 <0.72 <2.1 <1.6 <3.7 <3.8 <2.1 <4.7 <0.86 0.0 4.2 34 <0.71 <2.1 <1.6 <3.6 <3.8 <2.1 <4.6 <0.84 0.0 5.3 35 16 i 18 i 9 i <18 39 i 33 57 7.7 i 170.7 86

36 <6.3 <19 <14 <32 <34 <19 89 15 i 104.0 220 37 4.3 i 4.8 2.6 i <3.5 12.0 10 11 <0.81 42.1 3.7 i 38 24 16 6.6 <3.5 18.0 14 20 2.3 94.3 5.5 39 11 7.9 3.4 i <3.4 9.3 7.3 b 11 1.3 i 40.5 12 40 33 18 8.1 <3.5 28.0 19 94 20 212.0 260 41 27 17 7.1 <3.6 21.0 16 46 6.7 133.7 83 42 <0.64 <1.9 <1.4 <3.3 <3.4 <1.9 6.6 i 1.6 i 8.2 24 43 <0.66 <2 <1.5 <3.4 <3.5 <2 <4.3 <0.79 0.0 13 44 25 i <23 <17 82 i5 120 93 1100 150 1570 1900 45 13 6 4.1 i <3.3 41.0 35 250 49 394.0 290 46 <0.65 <2 <1.4 4.1 i <3.5 <2 21 1.8 26.9 31 47 1600 580 360 <210 1700 1500 14000 3200 22580 17000 48 2.8 i 2.4 i <1.6 <3.7 <3.8 2.6 i 19 4.3 i 31.1 38 49 4.5 i 2.7 i <1.8 15.0 6.6 i 5.2 i 38 10 82.0 100 50 <6.9 <21 <15 <35 <37 <21 390 100 490.0 1500 51A 5.7 2.9 i <1.6 <3.6 <3.7 <2.1 5.7 i <0.84 14.3 13 51B <0.72 <2.2 <1.6 <3.7 <3.8 <2.2 9.3 i 1.5 i 10.8 26 52A <52 <160 <120 <270 <280 220 i 7200 1300 8720 14000 52B <53 <160 <120 <270 <280 200 i 8000 1400 9600 15000 53 <3 <8.8 <6.5 23 i <16 <8.8 21 i 8.8 i 52.8 100 1 PAH abbreviations: Naph: Naphthalene; 2-Men: 2-methylnaphthalene; 1-Men: 1-methylnaphthalene; Acen: Acenaphthylene; Athen: Acenaphthene; Fluor: Fluorene; Phen: Phenanthrene; Anth” Anthracene; LMW: Light molecular weight PAH total; Flrant: Fluoranthene; S L&H: sum of LMW and HMW PAHs. 2 Bold faced font indicates that the concentration is equal to or exceeds the FDEP TEL. 3 Italicized underlined font indicates that the concentration is equal to or exceeds the FDEP PEL. 4 < indicates a nondetect showing that the result is below the MDL indicated by the number following. 5 i indicates that the reported value is between the laboratory method detection limit and the laboratory practical quantitation limit.

Table 17. PAH concentrations [ug/kg] in sediments, GBc series - continued. ID Py Ch BzaA BzbF BzkF BzaP InPy DzaA BzPe HMW S L&H Total TEL 153 108.0 74.8 88.8 6.2 655 1684 PEL 1398. 846.0 693.0 763.0 135.0 6676 16770 1 1.8 i <2.5 <1.1 <1.7 <2.3 <2 <1.6 <2.6 <1.6 1.8 1.8 1.8 2 7.1 6.4 4.6 9.7 7.1 6 5.5 <2.6 4.3 i 32.1 36.7 63.3 3 1.3 i <2.4 <1.1 <1.7 <2.3 <1.9 <1.6 <2.6 <1.6 1.3 1.3 1.3 4 28 39 25 84 79 58 30 7.9 18 190.9 216.3 427.3 5 <0.54 <2.3 <0.97 <1.6 <2.2 <1.9 <1.5 <2.5 <1.5 0 0 0 6 32 17 13 <1.8 <2.5 12 5.3 <2.8 4.5 i 102 109.9 119.7 7 34 26 14 24 21 15 13 <6 11 i 126 136 205 8 3 i <2.5 1.5 i <1.7 <2.4 <2 1.7 i <2.6 <1.6 7.4 7.4 9.1 9 21 16 11 <4.1 <5.6 16 8.8 i <6.2 <3.8 83 89.5 98.3 10 11 6.8 i 5.6 i <3.3 <4.6 6 i 3.2 i <5.1 <3.1 39.4 43.1 46.3 11 62 51 43 92 72 55 26 6.7 i 17 260.7 293.1 500.1 12 9.4 5.3 4.6 i <1.1 <0.89 4.3 i 2 i <0.7 1.6 i 33.1 38.5 42.1 13 11 6.9 4.9 i 8.5 7.8 5.6 i 3.3 i <0.85 <1.1 41.4 41.4 61 14 12 8.6 i 6.7 i 11 9.7 i 7.5 i 3.2 i <1.6 2.4 i 45.8 45.8 72.1 15 11 8.3 i 6.7 i 12 10 8.2 i 4.1 i <1.5 3 i 45.2 45.2 74.3 16 12 7.8 i 6.3 i 10 9.2 7.1 i 3.3 i <1.3 2.5 i 45.2 45.2 70.2 17 3.1 i 1.7 i 1.2 i <1.2 <0.91 <1.6 <1.3 <0.72 <0.91 9.4 9.4 9.4 18 <4.3 <0.8 <0.77 <0.68 <0.63 <0.66 <1.1 <0.85 <1.5 0 0 0 19 6.8 13 3.5 i <1.4 <1.2 <1.9 <1.5 <0.87 <1.2 29.6 29.6 29.6 20 <1.7 <8 <1.5 <2 <1.6 5.4 i <2.2 <1.3 <1.6 5.4 30.9 36.5 21 <0.99 <4.7 <0.86 4.5 i 3.8 i 2.9 i 1.8 i <0.72 <0.92 2.9 15.3 29.3 22 <1.2 <5.4 <0.99 7.1 6.4 5.7 3.7 i <0.83 3.1 i 5.7 30.3 57.4 23 <0.96 <4.6 <0.84 <1.1 <0.89 2 i <1.2 <0.7 <0.89 2 9.6 11.9 24 <0.95 <4.5 <0.83 <1.1 <0.88 <1.5 <1.2 <0.7 <0.88 0 2.1 4.2 25 <0.93 <4.4 <0.81 <1.1 <0.86 <1.5 <1.2 <0.68 <0.86 0 0 0 26 1.6 i 1.3 i <0.63 <1.1 <0.82 <1.4 <1.1 <0.64 <0.82 2.9 2.9 2.9 27 4.2 i 3 i 2.5 i 3.5 i 3.5 i 3.2 i 2.4 i <0.67 2 i 18 18 29.4 28 <0.67 <0.62 <0.64 <1.1 <0.84 <1.5 <1.2 <0.66 <0.84 0 0 0 29 7.6 5.3 4.2 i 5.3 3.7 i 3.3 i 1.8 i <0.66 <0.84 27.1 27.1 37.9 30 <0.67 <0.62 <0.64 <1.1 <0.83 <1.5 <1.2 <0.65 <0.83 0 0 0 31 1.1 <0.62 <0.65 <1.1 <0.84 <1.5 <1.2 <0.66 <0.84 1.1 1.1 1.1 32 22 19 14 23 19 18 17 <2.4 15 96 96 170 33 3.6 2.8 2.3 <1.2 <91 2.6 2.2 <0.72 2 15.5 15.5 19.7 34 4.4 i 3.6 i 2.7 i 4.1 i 4.2 i 3.4 i 2.9 i <0.71 2.5 i 19.4 19.4 33.1 35 77 57 41 75 60 57 46 5.6 i 39 323.6 494.3 723.3

36 200 150 100 170 170 140 110 14 i 93 824 928 1471 37 3 i 2.2 i 1.9 i <1.1 <0.86 2.2 i 1.7 i <0.68 1.4 i 13 55.1 60.8 38 3.9 i 2.2 i 2.1 i 2.7 i 2.2 i 2.4 1.7 i <0.69 1.9 i 16.1 110.4 125.5 39 8.6 4.8 1.7 i 3.9 i 3.2 i 1.7 i <1.2 <0.67 <0.85 28.8 69.3 87.1 40 270 450 290 700 470 630 270 90 280 1990 2202 3930.1 41 78 75 66 95 68 87 46 15 50 404 537.7 803.8 42 26 19 18 22 19 22 12 3.2 i 13 112.2 120.4 186.4 43 12 11 9.2 13 10 11 6.6 1.7 i 6.5 57.9 57.9 94 44 1700 1200 1000 1600 1100 1200 650 180 650 7180 8750 12750 45 230 150 150 190 120 150 77 22 76 992 1386 1853.1 46 29 16 11 17 14 14 8.6 1.8 8.6 102.8 129.7 177.9 47 14000 8400 7600 9000 8200 8000 2900 990 2700 55990 78570 101730 48 33 21 18 27 23 20 8.8 2.7 i 8.8 132.7 163.8 231.4 49 100 80 59 110 95 91 41 12 39 442 524 809 50 1600 1200 880 1400 1100 1000 430 140 400 6320 6810 10140 51A 12 9.9 7.1 14 11 8.8 5.1 <0.7 4.5 i 50.8 65.1 99.7 51B 24 17 14 23 19 15 6.8 2.1 i 6.5 98.1 108.9 164.2 52A 13000 8400 5600 7600 7400 6000 2500 940 2300 47940 56660 76460 52B 14000 8600 5800 8700 6900 6500 2500 880 2300 50780 60380 80780 53 120 79 65 140 97 94 44 12 i 44 470 522.8 847.8 1 PAH abbreviations: Naph: Naphthalene; 2-Men: 2-methylnaphthalene; 1-Men: 1-methylnaphthalene; Acen: Acenaphthylene; Athen: Acenaphthene; Fluor: Fluorene; Phen: Phenanthrene; Anth: Anthracene; LMW: Light molecular weight PAHs; Fant: Fluoranthene. 2 Bold faced font indicates that the concentration is equal to or exceeds the FDEP TEL. 3 Italicized underlined font indicates that the concentration is equal to or exceeds the FDEP PEL. 4 < indicates a nondetect showing that the result is below the MDL indicated by the number following. 5 i indicates that the reported value is between the laboratory method detection limit and the laboratory practical quantitation limit.

GBc-53

GF-15 GF-19

GF-23

Figure 23. Total PAH concentration in sediments. Triangles are GF series, circles are GBc series.

Figure 24. Sum of 13 LMW and HMW PAHs in sediments. Triangles are GF series, circles are GB series.

Table 18. PAH concentrations [ug/kg] in sediments, GF series. Sample Naph 2-Men 1-Men Acen Athen Fluor Phen Anth LMW Flrant Pyrene Chrys TEL 34.6 20.2 5.9 6.7 21.2 86.7 46.9 544.0 113.0 153.0 108.0 PEL 391.0 201.0 128.0 88.9 144.0 544.0 245.0 1442.0 1494.0 1398.0 846.0 GF-1 69 190 110 7.6 i 180 78 31 7.2 562.8 24 18 12 GF-2 350 900.0 500 57 800 330 210 48 2695.0 260 230 170 GF-3 210 520.0 260 44 i 410 170 340 65 1759.0 560 470 340 GF-4 310 1300.0 810 88 1800 830 400 70 4798.0 300 280 270 GF-5 130 83 41 i <42 78 i 37 i 71 i 27 i 426 200 180 150 GF-6 120 42 i 20 i <37 <39 <22 47 i 20 i 229 130 120 100 GF-7 200 180 100 <43 260 160 130 89 1019 150 130 110 GF-8 150 120 57 <39 66 i 39 i 72 i 63 510 85 72 51 GF-9 180 110 56 <38 87 i 59 88 i 40 i 564 57 51 33 I GF-10 17 8.5 4.1 i <3.3 <3.4 <1.9 <4.1 7.6 33.1 10 10 8.6 GF-11 120 41 22 <3.3 16 5.2 5.1 i 4.2 i 191.5 4.6 3.8 I 2.2 I GF-12 25 12 6.3 <3.3 5.4 i 3.4 i 6.7 i 5.6 58.1 5 4.1 I 2.3 I GF-13 290 120 62 57 i 130 81 120 110 908 220 190 160 GF-14 <11 <31 <23 <54 <56 <31 100 i 41 i 141 290 260 220 GF-15 <7.2 <22 <16 46 i <39 <22 77 i 37 i 160 250 220 170 GF-16 2.1 i <2 <1.5 5 i <3.6 2.1 i 9.4 4.7 23.3 19 17 20 GF-17 35 i 41 i 25 i 640 82 i 70 460 320 1648.0 1500.0 1400.0 1100 GF-18 42 110 65 <13 140 80 55 10 i 437 18 15 I 9.9 I GF-19 <0.79 7.4 4.6 i <4.0 14 8.4 16 3.1 i 48.9 25 23 19 GF-20 <3 9.6 i <6.4 <16 17 i <8.7 27 i <3.6 53.6 20 I 16 I <2.8 GF-21 <8.1 31 i <18 <42 66 i <24 58 i <9.7 155 <9.4 42 I <7.6 GF-22 <9.6 31 i <21 <49 56 i <29 82 i <12 169 140 130 100 GF-23 <11 36 i <23 <53 74 i 59 i 150 <13 319 180 180 130

See Table 17 for footnotes.

Table 18. PAH concentrations [ug/kg] for sediments, GF series - continued. Sample BzaA BzbF BzkF BzaP InPy DzaA BzPe HMW S L&H Total TEL 74.8 88.8 6.2 655 1684.0 PEL 693.0 763.0 135.0 6676 16770.0 GF-1 7.8 13 10 10 12 3.4 I 13 75.2 638 796 GF-2 130 190 140 160 160 32 120 982 3677 4787 GF-3 270 390 290 330 320 66 250 2036 3795 5305 GF-4 210 340 260 310 310 62 250 1432 6230 8200 GF-5 120 180 130 150 150 32 I 120 832 1258 1879 GF-6 77 130 100 110 120 24 I 97 561 790 1257 GF-7 77 110 95 100 98 17 I 85 584 1603 2091 GF-8 36 I 48 I 50 I 45 I 45 I <7.6 38 I 289 799 1037 GF-9 25 I 32 I 31 I 28 I 29 I <7.4 24 I 194 758 930 GF-10 5.7 5.7 5.5 3.7 I 3.2 I <0.81 2.5 I 38 71.1 92.1 GF-11 1.7 I <1 <0.81 1.4 I 1.3 I <0.64 1.1 I 13.7 205.2 229.6 GF-12 1.8 I 2 I 1.5 I 1.5 I 1.5 I <0.64 1.3 I 14.7 72.8 85.4 GF-13 120 160 160 160 160 31 I 130 881 1789 2461 GF-14 150 290 200 220 230 42 I 180 1182 1323 2223 GF-15 130 240 180 180 190 37 I 140 987 1147 1897 GF-16 14 23 18 19 16 3.1 I 12 92.1 115.4 184.4 GF-17 890 1300 1100 1300 1100 230 810 6420 8068 12403 GF-18 7.4 I <3.8 <3.1 9.3 I 8.8 <2.5 7.2 I 59.6 496.6 577.6 GF-19 12 21 17 15 13 4.1 I 15 98.1 147 217.6 GF-20 <2.9 <4.7 <3.8 <6.4 <5.1 <3 <3.8 36 89.6 89.6 GF-21 <7.9 <13 <11 <18 19 I <8.1 23 I 42 197 239 GF-22 80 120 97 94 69 28 I 81 572 741 1108 GF-23 98 150 140 120 91 26 I 110 734 1053 1544

See Table 17 for footnotes.

Studies by EPA and NOAA, reported in DeBusk et al. (2002), found a maximum PAH concentration of 1,285 ug/kg and a mean of 330 ug/kg in Bayou Grande sediments in the 1990’s. These concentrations are considerably lower than what was detected in the GF series in the current study where the maximum concentration for the sum of LMW and HMW PAHs was 8,068 ug/kg with a mean of 1525 ug/kg. The GBc series, without outliers GBc-47 and GBc-52, has a mean concentration of 471 ug/kg. It is tedious at best to compare data from different studies over time, but there is no evidence that the levels of PAHs have declined since those studies from the 1990’s and the variation in results is most likely a reflection of sampling patterns.

VII.1.2.2 PAHs in the water column PAHs were detected at trace level amounts in water samples GW-6, 7, and 8 (Table 19, Figure 20). The detected PAHs were the lighter weight molecular forms that included 2- methylnaphthalene and 1-methylnaphthalene that can originate from petroleum products. The same samples also showed detectable levels of total petroleum hydrocarbons (Table 15), which strongly suggests an origin from petroleum products for the PAHs.

VII.1.2.3 PAHs in vibracores Concentrations of most PAH species with the exception of naphthalenes decreased abruptly with depth with non-detects or low concentrations being observed in the lower levels. The highest total PAH concentrations were detected in level A (surface) from vibracore GV-4 (2017 ug/kg) from Woolsey Bayou near NAS (Table 20, Figure 21). Vibracore GV-4 had non-detects of PAHs in the B, C, and D levels. Core GV-1 had the second highest total PAH at level A and trace amounts in levels B and C. Vibracores GV-9 and 10 had low concentrations for level A and showed a less pronounced decrease in PAH concentration with depth. Overall the majority of detectable PAH concentrations occurred in surface sediments.

Heavy molecular weight PAHs were generally not detected at deeper sediment levels but naphthalenes, which are components of LMW PAHs, often were. The naphthalenes, and most other LMW PAHs in general, would not be expected to persist in the sediments for extended periods of time due to their volatility and degradability. Degradation of PAHs can occur in sediments aerobically and anaerobically (Hayes, 1999). Degradation of PAHs in deeper sediments is not readily predictable and depends upon the presence of suitable bacteria and organic substrates and sulfate, nitrate or Fe (III) reducing conditions. Sulfate reduction is typically the predominant anaerobic process under marine conditions (Stout et al. 2005). In a study by Coates et al. (1997) PAHs including naphthalene, phenanthrene, fluorene, fluoranthene and methylnaphthalene degraded anaerobically. All of the studied PAHs with the exception of fluoranthene were LMW PAHs. It is generally held that PAH biodegradation rates decline with an increase in the molecular mass that is related to increase in the quantity of aromatic rings (Elmendorf. et al., 1994). These observations indicate that the presence of LMW PAHs at the deeper levels suggests recent transport into the deeper sediments.

Table 19. PAHs and pentachlorophenol in water samples [ug/l]. PAH congener GW-1 GW-2 GW-3 GW-4 GW-5 GW-6 GW-7 GW-8 Naphthalene <0.0451 <0.045 <0.045 <0.045 <0.045 <0.045 <0.045 <0.045 2-Methylnaphthalene <0.037 <0.037 <0.037 <0.037 <0.037 0.088 i 0.095 i 0.14 1-Methylnaphthalene <0.036 <0.036 <0.036 <0.036 <0.036 0.051 i2 0.065 i 0.097 i Acenaphthylene <0.039 <0.039 <0.039 <0.039 <0.039 <0.039 <0.039 <0.039 Acenaphthene <0.039 <0.039 <0.039 <0.039 <0.039 0.098 ib3 0.3 b 0.19 b Fluorene <0.04 <0.04 <0.04 <0.04 <0.04 <0.04 <0.04 <0.04 Pentachlorophenol (PCP) <0.019 <0.019 <0.019 <0.019 <0.019 <0.019 <0.019 <0.019 Phenanthrene <0.037 <0.037 <0.037 <0.037 <0.037 0.11 0.33 <0.037 Anthracene <0.03 <0.03 <0.03 <0.03 <0.03 <0.03 <0.03 <0.03 Carbazole <0.025 <0.025 <0.025 <0.025 <0.025 <0.025 <0.025 <0.025 Fluoranthene <0.022 <0.022 <0.022 <0.022 <0.022 <0.022 <0.022 <0.022 Pyrene <0.021 <0.021 <0.021 <0.021 <0.021 <0.021 <0.021 <0.021 Benz(a)anthracene <0.014 <0.014 <0.014 <0.014 <0.014 <0.014 <0.014 <0.014 Chrysene <0.011 <0.011 <0.011 <0.011 <0.011 <0.011 <0.011 <0.011 Benzo(b)fluoranthene <0.015 <0.015 <0.015 <0.015 <0.015 <0.015 <0.015 <0.015 Benzo(k)fluoranthene <0.015 <0.015 <0.015 <0.015 <0.015 <0.015 <0.015 <0.015 Benzo(a)pyrene <0.014 <0.014 <0.014 <0.014 <0.014 <0.014 <0.014 <0.014 Indeno(1,2,3-cd)pyrene <0.016 <0.016 <0.016 <0.016 <0.016 <0.016 <0.016 <0.016 Dibenz(a,h)anthracene <0.015 <0.015 <0.015 <0.015 <0.015 <0.015 <0.015 <0.015 Benzo(g,h,i)perylene <0.015 <0.015 <0.015 <0.015 <0.015 <0.015 <0.015 <0.015

1 < indicates a nondetect showing that the result is below the MDL indicated by the number following. 2 i indicates that the reported value is between the laboratory method detection limit and the laboratory practical quantitation limit. 3 b indicates that the analyte was present in method blank.

Table 20. Total PAHs in vibracores by depth level [ug/kg]. level A level B level C level D Core surface 1 m depth 2 m depth 3 m depth GV-1 1426 2.4 15.1 NS1 GV-2 13.3 0 0 NS GV-4 2017 0 0 0 GV-5 4.2 2.1 0 NS GV-6 228 0 8.3 NS GV-7 102 0 0 NS GV-8 7.4 0 NS NS GV-9 69.2 7.4 2.1 16.7 GV-10 53.7 4.8 34.6 NS GV-11 912 0 0 39 1 NS indicates that the indicated level was not reached in that core.

Seven of the ten vibracores showed the presence of naphthalenes (naphthalene, 2- methylnaphthalene, 1-methylnaphthalene). These naphthalenes were the most common PAHs in the deeper sediments (Table 21). Some other PAHs were also detected in deeper sediments such as benzo(a)pyrene, a carcinogen detected below its action level (100 ug/kg FDEP Clean up Level) in sample GV-10C. Sample GV-1 was taken near sample GBc-52 that had the second highest concentrations of total surface PAHs. Vibracore GV-1 does have trace detections of 1- methylnaphthalene, 2-methylnaphthalene, and acenaphthene that are light molecular weight PAHs at the 2 meter level as well as several heavy molecular weight PAHs. The detections of the heavier PAHs included pyrene, benzo(a)pyrene, indeno(1,2,3-cd)pyrene, benzo(g,h,i)perylene. The presence of these heavier PAHs is most likely not due to transport by a contaminated aquifer because of their lower solubilities in water. It is possible that the heavier PAHs were laid down at the same time as the sediments were laid down or were introduced at a later date perhaps by driving a tarred pole or other structure into the sediments. In any case the detection of the heavier PAHs is at trace levels of 2.1 to 4.1 ug/kg for this sample and the heavier PAHs were only found to be present in deeper sediments for one other vibracore sample (GV-10C) that is also in an area that could have had a dock made of treated wood at one time.

Cores GV-2, 4, 7 and 8 had no detections of PAHs at deeper sediment levels. Core GV-6 taken at the head of a small dredged embayment densely surrounded by residences showed detections of 1-methylnaphthalene, 2-methylnaphthalene at 2 meters depth. This could represent sediment contamination by contaminated groundwater coming from elsewhere than NAS since this site appears very distant from NAS influence. The other cores with naphthalenes present at subsurface levels could be under the influence of aquifer transport of these contaminants. Vibracores GV-9, 10 and 11 that are in the main part of the Bayou all showed naphthalene detection at their deepest sediment levels. The observation of naphthalene at levels ranging from 1-3 m in six out of ten vibracores suggests that there is general naphthalene contamination of the deeper bayou sediments. However, the data does not provide conclusive evidence that any specific site is solely or in part the cause of it. It is just as likely that documented sites on NAS or undocumented sites on the Warrington side of the Bayou are responsible. It is suggested that the naphthalene observed in the bayou sediments is likely due to transport by contaminated groundwater although surface transport mechanisms cannot be completely ruled out.

76 Grande rpt working 77 of 175 8/25/2008 Table 21. Specific PAHs in Bayou Grande vibracores [ug/kg]. PAH GV-1 GV-2 GV-4 level A level B level C level A level B level C level A level B level C level D Naphthalene <7.7 <0.75 <0.68 <0.66 <0.63 <0.63 <3.5 <0.73 <0.63 <0.62 2-Methylnaphthalene <23 2.4i 2.2i 2.9i <1.9 <1.9 <11 <2.2 <1.9 <1.9 1-Methylnaphthalene <17 <1.7 1.8i 2.1i <1.4 <1.4 <7.6 <1.6 <1.4 <1.4 Acenaphthylene <40 <3.9 <3.5 <3.4 <3.2 <3.2 26i <3.8 <3.2 <3.2 Acenaphthene <41 <4 <3.6 8.3i <3.3 <3.3 30i <3.9 <3.3 <3.3 Fluorene <23 <2.3 <2 <2 <1.9 <1.9 <11 <2.2 <1.9 <1.9 Phenanthrene 62i <4.9 <4.4 <4.3 <4.1 <4.1 160 <4.8 <4.1 <4 Anthracene 24i <0.9 <0.81 <0.79 <0.75 <0.75 33 <0.88 <0.75 <0.74 Fluoranthene 190 <0.87 <0.78 <0.76 <0.72 <0.72 290 <0.85 <0.72 <0.72 Pyrene 200 <0.77 4.1i <0.67 <0.64 <0.64 270 <0.75 <0.64 <0.63 Chrysene 140 <0.71 <0.64 <0.62 <0.59 <0.59 190 <0.69 <0.59 <0.58 Benz(a)anthracene 100j <0.74 <0.66 <0.65 <0.61 <0.61 150 <0.72 <0.61 <0.61 Benzo(b)fluoranthene 170j <1.2 <1.1 <1.1 <0.98 <0.98 210 <1.2 <0.98 <0.97 Benzo(k)fluoranthene 150j <0.96 <0.86 <0.84 <0.8 <0.8 170 <0.93 <0.8 <0.79 Benzo(a)pyrene 130j <1.7 2.4i <1.5 <1.4 <1.4 180 <1.6 <1.4 <1.4 Indeno(1,2,3-cd)pyrene 120j <1.3 2.1i <1.2 <1.1 <1.1 130 <1.3 <1.1 <1.1 Dibenz(a,h)anthracene <7.7 <0.75 <0.68 <0.66 <0.63 <0.63 38 <0.73 <0.63 <0.62 Benzo(g,h,i)perylene 140j <0.96 2.5i <0.84 <0.8 <0.8 140 <0.93 <0.8 <0.79

77 Grande rpt working 78 of 175 8/25/2008 Table 21. Specific PAHs in Bayou Grande vibracores - continued. PAH GV-5 GV-6 GV-7 level A level B level C level A level B level C level A level B level C Naphthalene <0.66 <0.64 <0.64 <5.4 <0.76 <0.63 <4.9 <0.61 <0.65 2-Methylnaphthalene <2 2.1i <1.9 <16 <2.3 5.4 <15 <1.8 <1.9 1-Methylnaphthalene <1.5 <1.4 <1.4 <12 <1.7 2.9i <11 <1.3 <1.4 Acenaphthylene <3.4 <3.3 <3.3 <28 <3.9 <3.2 <25 <3.1 <3.3 Acenaphthene <3.5 <3.4 <3.4 <29 <4 <3.4 <26 <3.2 <3.5 Fluorene <2 <1.9 <1.9 <16 <2.3 <1.9 <15 <1.8 <1.9 Phenanthrene <4.3 <4.2 <4.2 <35 <4.9 <4.1 45i <3.9 <4.2 Anthracene <0.79 <0.77 <0.77 <6.5 <0.9 <0.75 <5.8 <0.73 <0.77 Fluoranthene <0.77 <0.74 <0.74 69 <0.88 <0.73 <5.7 <0.7 <0.75 Pyrene <0.68 <0.66 <0.65 81 <0.77 <0.64 57 <0.62 <0.66 Chrysene <0.62 <0.61 <0.6 <5.1 <0.71 <0.59 <4.6 <0.57 <0.61 Benz(a)anthracene <0.65 <0.63 <0.63 <5.3 <0.74 <0.62 <4.8 <0.6 <0.63 Benzo(b)fluoranthene <1.1 <1.1 <1 <8.5 <1.2 <0.99 <7.6 <0.95 <1.1 Benzo(k)fluoranthene <0.84 <0.82 <0.82 <6.9 <0.96 <0.8 <6.2 <0.77 <0.82 Benzo(a)pyrene 2.2i <1.4 <1.4 <12 <1.7 <1.4 <11 <1.3 <1.4 Indeno(1,2,3-cd)pyrene 2i <1.1 <1.1 31i <1.3 <1.1 <8.3 <1.1 <1.1 Dibenz(a,h)anthracene <0.66 <0.64 <0.64 <5.4 <0.76 <0.63 <4.9 <0.61 <0.65 Benzo(g,h,i)perylene 2.6i <0.82 <0.82 47 <0.96 <0.8 <6.2 <0.77 <0.82

78 Grande rpt working 79 of 175 8/25/2008 Table 21. Specific PAHs in Bayou Grande vibracores - continued. PAH GV-8 GV-9 GV-10 GV-11 level A Level B level A level B level C level D level A level B level C level A level B level C level D Naphthalene <0.66 <1.22 111 <0.63 <0.63 <0.64 <0.63 <0.64 <0.83 <8.1 <4.4 <4.7 24i3 2-Methylnaphthalene <2 <3.3 23 4.8 2.1i 7.4 7.4i 3.2i 2.6i <24 <13 <14 15i 1-Methylnaphthalene <1.5 <2.4 13 2.6 i <1.4 3.7i 3.8i 1.6i <1.8 <18 <9.5 <11 <8.5 Acenaphthylene <3.4 <5.7 <3.2 <3.2 <3.3 <3.3 <3.2 <3.3 <4.3 <42 <23 <24 <21 Acenaphthene <3.5 <5.9 <3.3 <3.3 <3.4 <3.4 <3.4 <3.4 <4.4 <43 <24 <25 <21 Fluorene <2 <3.3 <1.9 <1.9 <1.9 <1.9 <1.9 <1.9 <2.5 <24 <13 <14 <12 Phenanthrene <4.3 <7.2 4.5i <4.1 <4.1 5.6i 5.1i <4.2 <5.4 <53 <29 <31 <26 Anthracene <0.79 <1.4 <0.75 <0.75 <0.76 <0.77 <0.75 <0.77 <0.99 <9.7 <5.3 <5.6 <4.7 Fluoranthene <0.77 <1.3 <0.73 <0.72 <0.73 <0.75 7.5 <0.74 <0.96 120 <5.1 <5.4 <4.6 Pyrene 5.5 <1.2 5.9 <0.64 <0.65 <0.66 7.8 <0.66 <0.85 150 <4.5 <4.8 <4.1 Chrysene <0.63 <1.1 <0.59 <0.59 <0.6 <0.61 4.8 <0.61 <0.78 98 <4.2 <4.4 <3.7 Benz(a)anthracene <0.65 <1.1 <0.62 <0.61 <0.62 <0.63 4.1i <0.63 <0.81 79 <4.3 <4.6 <3.9 Benzo(b)fluoranthene <1.1 <1.8 <0.98 <0.98 <0.99 <1.1 <0.99 <1.1 <1.3 120 <6.9 <7.3 <6.2 Benzo(k)fluoranthene <0.85 <1.5 <0.8 <0.8 <0.81 <0.82 <0.8 <0.82 <1.1 86 <5.6 <6 <5.1 Benzo(a)pyrene 1.9i <2.4 4i <1.4 <1.4 <1.4 4.4 <1.4 32 100 <9.5 <11 <8.5 Indeno(1,2,3-cd)pyrene <1.2 <1.9 3.4i <1.1 <1.1 <1.1 4i <1.1 <1.5 70 <7.5 <8 <6.7 Dibenz(a,h)anthracene <0.66 <1.2 <0.63 <0.63 <0.63 <0.64 <0.63 <0.64 <0.83 20i <4.4 <4.7 <4 Benzo(g,h,i)perylene <0.85 <1.5 4.4 <0.8 <0.81 <0.82 4.8 <0.82 <1.1 89 <5.6 <6 <5.1 1 Bold faced font indicates that the concentration is equal to or exceeds the FDEP TEL. 2 < indicates a nondetect showing that the result is below the MDL indicated by the number following. 3 i indicates that the reported value is between the laboratory method detection limit and the laboratory practical quantitation limit.

VII.1.2.4 Origins of PAHs in Bayou Grande Surface grabs GF-1 to 23, which were collected away from the shoreline in the various channel areas of Bayou Grande and its embayments, have detections of naphthalenes (Figure 25, Table 18). Sample GF-2 (Yacht Basin), GF-3 (Woolsey Bayou), and GF-4 (Bayou Grande near NAS) exceed the FDEP PEL for 2-methylnaphthalene that can be present in petroleum products spills. Naphthalene and 1-methylnaphthalene were also found at elevated concentrations in these samples (Table 22, Table 23). The fact that these samples originated near areas that are associated with aviation activities strongly suggests a contribution from petroleum fuels to the sediment PAHs (Table 22). Sample GBc-47, also collected in the Yacht Basin but closer to this likely source than GF-2, has significant concentrations of naphthalenes. GBc-48 that is just a short distance away has lower concentrations of PAHs that do not exceed SQAG. It could be postulated that the high naphthalene concentrations in GBc-47 might occur as a predicted component of such a massive concentration of PAHs. However, in the sample with the next highest concentration that is from Navy Point Bayou (GBc-52B, 80,780 ug/kg) no naphthalenes were detected. This demonstrates that the higher concentrations of naphthalenes in GBc-47 are not solely a function of its overall PAH content since GBc-52 had no detections. For the remaining GBc series that were mostly located near shorelines, the naphthalene PAH species range from non-detect to very low detected concentrations, indicating that samples taken close to the shoreline but away from the more intense NAS activities do not have significant naphthalenes. In samples from the GF series that were mostly located in the basin or deeper channels of the bayou and its embayments, there was detection of 2-methylnaphthalene in 20 out 23 samples suggesting that there is a general distribution of naphthalenes in the surface sediments from the deeper regions of Bayou Grande. The EnSafe (2003, 2004) studies of PAHs in Bayou Grande show that naphthalenes were also present near NAS (Table 5). An examination of Figure 25 shows that total naphthalenes concentrations are highest in embayments and shorelines adjacent to NAS but that samples from Bayou Grande’s channel more consistently had detectable concentrations of naphthalenes than was the case for the GBc series of samples that are located mostly near shorelines. Tables 22 and 23 show that on the basis of percent composition, naphthalene represents a greater percentage for samples in the GF series (mean 22%) than for those in the GBc series (mean 8%). This suggests general contamination of the surface sediments in the Bayou’s main channel by naphthalenes.

Figure 25. Total naphthalenes in surface sediments. Triangles are GF series, circles are GB series.

If the PAHs are derived from petroleum releases, on land one would expect the total naphthalene ratio and concentration to have higher values near outfalls and to coincide with the highest total PAH concentrations. The highest concentrations of total naphthalenes and total PAH were observed in the Yacht Club Basin embayment, but the ratios for naphthalenes versus total PAH were substantially lower than what was observed in many areas of the Bayou’s channel (Tables 22 and 23). Naphthalenes are characteristically more volatile and soluble than the heavier PAHs. The greater solubility allows them to be transported in underground plumes to greater distances from land based release sites than is the case for many other species of PAHs under similar conditions. A spill of petroleum based aviation fuels into the aquifer would be expected to transport the naphthalenes further and in greater amounts than the other PAHs. The high proportion of naphthalenes in the deeper channel areas suggests groundwater transport rather than stormwater transport as the most probable mechanism for their origin.

Table 22. Naphthalene concentrations [ug/kg] in sediments, GF series. Sample ID Tot naphs1 % naphs/Tot PAHs2 Tot PAH3 GF-1 369 46.4 796 GF-2 1750 36.6 4787 GF-3 990 18.7 5305 GF-4 2420 29.5 8200 GF-5 254 13.5 1879 GF-6 182 14.5 1257 GF-7 480 23.0 2091 GF-8 327 31.5 1037 GF-9 346 37.2 930 GF-10 29.6 32.1 92.1 GF-11 183 79.7 229.6 GF-12 43.3 50.7 85.4 GF-13 472 19.2 2461 GF-14 0 0.0 2223 GF-15 0 0.0 1897 GF-16 2.1 1.1 184.4 GF-17 101 0.8 12403 GF-18 217 37.6 577.6 GF-19 12 5.5 217.6 GF-20 9.6 10.7 89.6 GF-21 31 13.0 239 GF-22 31 2.8 1108 GF-23 36 2.3 1544 1 Tot naphs is the sum of naphthalene, 2-methylnaphthalene, and 1-methylnaphthalene in ug/kg. 2 % Naphs/Tot PAH is the percent of Total naphthalenes relative to the total PAHs. 3 Tot PAH is the sum of all detected PAHs (18) ug/kg

PAHs can have multiple origins with oil spills and combustion products being among the most important sources in typical urban environments. Ratios based on concentrations of specific PAHs present within a sediment sample have been employed to obtain evidence that suggests the probable origins of PAH mixtures (Rostad and Pereira, 1987; Yunker et al., 2002). According to Rostad and Pereira (1987) typical coastal sediment PAHs are not of coal tar origin and have less phenanthrene than fluoranthene or pyrene. In the Rostad and Pereira (1987) system normalized ratios for fluoranthene or pyrene to phenanthrene of less than 100 suggest a coal tar or creosote origin. Yunker et al. (2002) listed four PAH ratio calculations for parent (non-alkylated) PAH compounds. Three of the four ratio calculations were applied to this study (Table 24). The ratios can in principle be correlated with one of four sources: petroleum release; combustion of petroleum products; combustion of grass, wood, and/or coal; and creosote origin. It was previously suggested that in some regions of Bayou Grande PAHs containing significant amounts of naphthalenes may be from petroleum fuel releases. Examination of these PAH ratios can suggest the origin of some of the remaining 15 detected PAHs.

Table 23. Naphthalene concentrations [ug/kg] in sediments, GBc series. Sample % naphs/ % naphs/ Tot naphs1 Tot PAH3 Sample ID Tot naphs Tot PAH ID Tot PAH2 Tot PAH GBc-1 0 0 1.8 GBc-29 0 0.0 37.9 GBc-2 0 0 63.3 GBc-30 0 0.0 0 GBc-3 0 0 1.3 GBc-31 0 0.0 1.1 GBc-4 0 0 427.3 GBc-32 0 0.0 170 GBc-5 0 0 0 GBc-33 0 0.0 19.7 GBc-6 0 0 119.7 GBc-34 0 0.0 33.1 GBc-7 0 0 205 GBc-35 43 5.9 723.3 GBc-8 0 0 9.1 GBc-36 0 0.0 1471 GBc-9 0 0 98.3 GBc-37 11.7 19.2 60.8 GBc-10 0 0 46.3 GBc-38 46.6 37.1 125.5 GBc-11 0 0 500.1 GBc-39 22.3 25.6 87.1 GBc-12 0 0 42.1 GBc-40 59.1 1.5 3930.1 GBc-13 0 0 61 GBc-41 51.1 6.4 803.8 GBc-14 0 0 72.1 GBc-42 0 0.0 186.4 GBc-15 0 0 74.3 GBc-43 0 0.0 94 GBc-16 0 0 70.2 GBc-44 25 0.2 12750 GBc-17 0 0 9.4 GBc-45 23.1 1.2 1853.1 GBc-18 0 0 0 GBc-46 0 0.0 177.9 GBc-19 0 0 29.6 GBc-47 2540 2.5 101730 GBc-20 24.4 66.8 36.5 GBc-48 5.2 2.2 231.4 GBc-21 13.3 45.4 29.3 GBc-49 7.2 0.9 809 GBc-22 26.6 46.3 57.4 GBc-50 0 0.0 10140 GBc-23 8.1 68.1 11.9 GBc-51A 8.6 8.6 99.7 GBc-24 4.2 100 4.2 GBc-51B 0 0.0 164.2 GBc-25 0 0 0 GBc-52A 0 0.0 76460 GBc-26 0 0 2.9 GBc-52B 0 0.0 80780 GBc-27 0 0 29.4 GBc-53 0 0.0 847.8 GBc-28 0 0 0 1 Tot naphs is the sum of naphthalene, 2-methylnaphthalene, and 1-methylnaphthalene in ug/kg. 2 % Naphs/Tot PAH is the percent of Total naphthalenes relative to the total PAHs. 3 Tot PAH is the sum of all detected PAHs (18) ug/kg

Table 24. PAH origin indicator ratios, Yunker et al. (2002). Petroleum Vehicle, crude oil Combustion grass, Creosoted PAH Ratio release combustion wood, coal wood pilings An/(Pn+An)1 <0.10 >0.10 >0.10 >0.18 Fl/(Fl+Py)2 <0.40 0.40-0.50 >0.50 >0.62 IP/(IP+Bghi)3 <0.20 0.20 to 0.50 >0.50 >0.62 1 An/(Pn+An)= Ratio of Anthracene/ Anthracene+ Phenanthrene 2 Fl/(Fl+Py) = Ratio of Fluoranthene/ Fluoranthene+ Pyrene 3 IP/(IP+Bghi) = Ratio of Indeno(1,2,3-c,d)Pyrene/ Indeno(1,2,3-c,d)Pyrene + Benzo(g,h,i) Perylene

The Rostad and Pereira (1987) method indicates the presence of PAHs of a coal tar origin when ratios of 100 or less are present. Ten out of 23 of the GF series have such a ratio of 100 or less for pyrene/phenanthrene. Ratios for several other samples are just above 100 (Table 25). The fluoranthene/phenanthrene ratios have comparable values (Table 25). Results for sediments for the GBc series also have some values of less than 100, but not to the same extent (Table 26). These high ratios suggest coal tar origins for the PAHs in Bayou Grande. It appears unlikely that there are any substantial creosote sources other than some treated wood. A more likely source would be coal tars used for other purposes or improperly disposed. In Bayou Grande there is a record of coal tar releases at the NAS facility (EnSafe, 2003; 2004).

For the GF series of samples the An/(Pn+An) ratio from Yunker et al. suggests a coal tar origin for 14 out of 23 of the samples (Table 25). The Fl/(Fl+Py) and IP/(IP+Bghi) ratios, however, overwhelmingly suggest combustion of grass, wood, and/or coal products as the origin for these PAHs in the channel of the Bayou. For the GBc series from the shorelines and the more distal embayment reaches the Rostad and Pereira ratios and An/(Pn+An) ratios show values indicating a mixed origin with combustion and coal tar origins for the sediment PAHs (Table 26). The IP/IP+Bghi ratios indicate a brush, forest, and/or coal combustion origin for the PAHs with some ratios also indicating combustion of petroleum products.

In summary, the PAHs in Bayou Grande sediments exhibit characteristics that suggest multiple origins but not petroleum spills. The total PAH composition, however, has a significant naphthalene content that suggests an input from petroleum fuels. It appears that since the forensic ratios do not suggest a fuel origin that petroleum spills are not primarily responsible for the bulk of PAHs in the Bayou. There is a strong likelihood of an NAS contribution for some of the PAHs found in bayou sediments since it was stated in the Superfund Program Proposed Plan for Bayou Grande (EnSafe, 2003; 2004) that the NAS wastes for semivolatile organic compounds included common components of asphalt, coal tar, and jet and diesel fuels. Brush fires on the shores and combustion from internal combustion engines are also likely sources. It also appears that the northern shores have also made a contribution, as demonstrated by the high PAH values in Navy Point Bayou. What is unique about these data from the channel and NAS is that there are high concentrations of naphthalenes without the characteristic ratios that indicate a spill of petroleum being present. Aviation fuels do certainly contain other PAHs than naphthalenes, as discussed in a recent study of degradation of aviation fuels in soils (Oleszczuk and Baran, 2003). The important mechanisms in play may be solubility in that the more soluble two-ring naphthalenes are more transportable by groundwater than the heavier PAHs. As a result, the less soluble PAHs may remain closer to the land based release site without undergoing significant transport by ground water. Transport of the lighter PAHs has been observed at the Escambia Treating Site in Pensacola where LMW PAHs are transported in groundwater to points very distant from the release point (CDM Federal Programs Corporation, 2002).

Table 25. PAH ratios in sediments, GF series. 1 2 3 4 ID Fl*100/Ph Py*100/Ph An/(Pn+An) Fl/(Fl+Py) IP/(IP+Bghi) GF-1 77 58 0.19 0.57 0.48 GF-2 124 110 0.19 0.53 0.57 GF-3 165 138 0.16 0.54 0.56 GF-4 75 70 0.15 0.52 0.55 GF-5 282 254 0.28 0.53 0.56 GF-6 277 255 0.30 0.52 0.55 GF-7 115 100 0.41 0.54 0.54 GF-8 118 100 0.47 0.54 0.54 GF-9 65 58 0.31 0.53 0.55 GF-10 No Ratio No Ratio No Ratio 0.50 0.56 GF-11 90 75 0.45 0.55 0.54 GF-12 75 61 0.46 0.55 0.54 GF-13 183 158 0.48 0.54 0.55 GF-14 290 260 0.29 0.53 0.56 GF-15 325 286 0.32 0.53 0.58 GF-16 202 181 0.33 0.53 0.57 GF-17 326 304 0.41 0.52 0.58 GF-18 33 27 0.15 0.55 0.55 GF-19 156 144 0.16 0.52 0.46 GF-20 74 59 No Ratio 0.56 No Ratio GF-21 No Ratio 72 No Ratio No Ratio 0.45 GF-22 171 159 No Ratio 0.52 0.46 GF-23 120 120 No Ratio 0.50 0.45 1: Rostad/Pereira Ratios are based on calculating Phenanthrene to 100 and calculating the relative percent difference to Fluoranthene and Pyrene, i.e. (100/μg Phenanthrene) x μg of Phenanthrene, Fluoranthene, or Pyrene 2: An/(Pn+An)= Ratio of Anthracene/ Anthracene+ Phenanthrene 3: Fl/(Fl+Py) = Ratio of Fluoranthene/ Fluoranthene+ Pyrene 4: IP/(IP+Bghi) = Ratio of Indeno(1,2,3-c,d)Pyrene/ Indeno(1,2,3-c,d)Pyrene+Benzo(g,h,i) Perylene

Table 26. PAH ratios in sediments, GBc series. Sample Fl 100/Ph1 Py 100/Ph An/(Pn+An) 2 Fl/(Fl+Py) 3 IP/(IP+Bghi) 4 ID * * GBC-1 No Ratio No Ratio No Ratio No Ratio No Ratio GBC-2 364 323 No Ratio 0.53 0.56 GBC-3 No Ratio No Ratio No Ratio No Ratio No Ratio GBC-4 471 400 0.55 0.54 0.63 GBC-5 No Ratio No Ratio No Ratio No Ratio No Ratio GBC-6 737 842 0.31 0.47 0.54 GBC-7 370 340 No Ratio 0.52 0.54 GBC-8 No Ratio No Ratio No Ratio 0.49 No Ratio GBC-9 475 525 0.38 0.48 No Ratio GBC-10 270 297 No Ratio 0.48 No Ratio GBC-11 672 969 0.63 0.41 0.60 GBC-12 175.93 174.07 No Ratio 0.50 0.56 GBC-13 No Ratio No Ratio No Ratio 0.54 No Ratio GBC-14 No Ratio No Ratio No Ratio 0.48 0.57 GBC-15 No Ratio No Ratio No Ratio 0.50 0.58 GBC-16 No Ratio No Ratio No Ratio 0.50 0.57 GBC-17 No Ratio No Ratio No Ratio 0.52 No Ratio GBC-18 No Ratio No Ratio No Ratio No Ratio No Ratio GBC-19 No Ratio No Ratio No Ratio 0.48 No Ratio GBC-20 No Ratio No Ratio No Ratio No Ratio No Ratio GBC-21 No Ratio No Ratio No Ratio No Ratio No Ratio GBC-22 No Ratio No Ratio No Ratio No Ratio 0.54 GBC-23 No Ratio No Ratio No Ratio No Ratio No Ratio GBC-24 No Ratio No Ratio No Ratio No Ratio No Ratio GBC-25 No Ratio No Ratio No Ratio No Ratio No Ratio GBc-26 No Ratio No Ratio No Ratio No Ratio No Ratio GBc-27 No Ratio No Ratio No Ratio 0.55 0.55 GBc-28 No Ratio No Ratio No Ratio No Ratio No Ratio GBc-29 No Ratio No Ratio No Ratio 0.47 No Ratio GBc-30 No Ratio No Ratio No Ratio No Ratio No Ratio GBc-31 No Ratio No Ratio No Ratio No Ratio No Ratio GBc-32 No Ratio No Ratio No Ratio 0.51 0.53 GBc-33 No Ratio No Ratio No Ratio 0.54 0.52 GBc-34 No Ratio No Ratio No Ratio 0.55 0.54 GBc-35 151 135 0.12 0.53 0.54 GBc-36 247 225 0.14 0.52 0.54 GBc-37 34 27 No Ratio 0.55 0.55 GBC-38 28 20 0.10 0.59 0.47 GBC-39 109 78 0.11 0.58 No Ratio GBC-40 277 287 0.18 0.49 0.49 GBC-41 180 170 0.13 0.52 0.48 GBC-42 364 394 0.20 0.48 0.48 GBC-43 No Ratio No Ratio No Ratio 0.52 0.50 GBC-44 173 155 0.12 0.53 0.50 GBC-45 116 92 0.16 0.56 0.50 GBC-46 148 138 0.08 0.52 0.50

Table 26. PAH ratios in sediments, GBc series - continued. Sample Fl 100/Ph1 Py 100/Ph An/(Pn+An) 2 Fl/(Fl+Py) 3 IP/(IP+Bghi) 4 ID * * GBC-47 121 100 0.19 0.55 0.52 GBC-48 200 174 No Ratio 0.54 0.50 GBC-49 263 263 0.21 0.50 0.51 GBC-50 385 410 0.20 0.48 0.52 GBC-51A 228 211 No Ratio 0.52 0.53 GBC-51B 280 258 0.14 0.52 0.51 GBC-52A 194 181 0.15 0.52 0.52 GBC-52B 188 175 0.15 0.52 0.52 GBC-53 476 571 0.30 0.45 0.50 1: Rostad/Pereira Ratios are based on calculating Phenanthrene to 100 and calculating the relative percent difference to Fluoranthene and Pyrene, i.e. (100/μg Phenanthrene) x μg of Phenanthrene, Fluoranthene, or Pyrene 2: An/(Pn+An)= Ratio of Anthracene/ Anthracene+ Phenanthrene 3: Fl/(Fl+Py) = Ratio of Fluoranthene/ Fluoranthene+ Pyrene 4: IP/(IP+Bghi) = Ratio of Indeno(1,2,3-c,d)Pyrene/ Indeno(1,2,3-c,d)Pyrene+Benzo(g,h,i) Perylene

The presence of naphthalenes suggests fuel or solvent releases. However, there is little evidence of any recent fuel releases via the surface into Bayou Grande. Intrusion and transport of contaminated ground water into the Bayou is possible and NAS is a likely source of such contamination. One potential naphthalene source is Operating Unit 1 (OU1) that is a former 85 acre sanitary landfill near the Bayou Grande shoreline. In the Final Five-Year Review of NAS (Tetra Tech, 2003) it was stated that there were problems with OU1. The USEPA comments mention what appear to be deficiencies of its treatment system and natural attenuation augmentation for OU1. Wastes consisted of ketones, PCBs and transformer oil-soaked rags, paint chips, paint sludge, asbestos, and garbage, and a tar pit was located nearby. The shallow and intermediate groundwater analytes included several solvents, metals, 2-methylnaphthalene, naphthalene, and other compounds. There exists the possibility that the naphthalenes in this inactive landfill or perhaps from other release sites on the NAS facilities could account at least in part for the naphthalenes observed in bayou sediments. There also exists the possibility that the naphthalenes in the groundwater might serve to mediate the transport of other more hydrophobic molecules such as PCBs that are also present at this site.

VII.1.2.5 Pentachlorophenol (PCP) There were no detections of pentachlorophenol and this is in agreement with the lack of any known releases or other sources of PCP.

VII.2 Dioxins/furans and PCBs

VII.2.1 Dioxin/furan and PCB TEQ The relative toxicity of dioxins/furans and dioxin-like PCBs is determined by toxic equivalence (TEQ). The total dioxin TEQ value expresses the toxicity as if the mixture were pure TCDD (2,3,7,8-tetrachlorodibenzo-p-dioxin), which exhibits the highest degree of toxicity. The TEQ approach and current toxicity equivalency factors (TEFs) (Table 13) have been adopted internationally as the most appropriate way to estimate the potential health risks of these dioxins/furans. In June 2005, a World Health Organization (WHO)-International Program meeting was held in Geneva during which the TEFs for dioxin-like compounds, including some polychlorinated biphenyls (PCBs), were reevaluated. It is these reevaluated TEFs that were used in the current study. Decisions about the TEF values were made in the WHO meeting based on a combination of the unweighted relative effect potency distributions from this database, expert judgment, and point estimates. Previous TEFs were assigned in increments of 0.01, 0.05, 0.1, etc., but for this reevaluation it was decided to use half order of magnitude increments on a logarithmic scale of 0.03, 0.1, 0.3, etc. Changes were decided by the expert panel for 2,3,4,7,8- pentachlorodibenzofuran (PeCDF) (TEF = 0.3), 1,2,3,7,8-pentachlorodibenzofuran (PeCDF) (TEF = 0.03), octachlorodibenzo-p-dioxin (OCCD) and octachlorodibenzofuran (OCCF) (TEFs = 0.0003), 3,4,4',5-tetrachlorbiphenyl (PCB 81) (TEF = 0.0003), 3,3',4,4',5,5'- hexachlorobiphenyl (PCB 169) (TEF = 0.03), and a single TEF value (0.00003) for all relevant mono-ortho–substituted PCBs. “Additivity”, an important prerequisite of the TEF concept was again confirmed by results from in vivo mixture studies. Certain individual and groups of compounds were identified for possible future inclusion in the TEF concept, including 3,4,4'- TCB (PCB 37), polybrominated dibenzo-p-dioxins and dibenzofurans, mixed polyhalogenated dibenzo-p-dioxins and dibenzofurans, polyhalogenated naphthalenes, and polybrominated biphenyls. There are also arguments for inclusion of polybrominated diphenyl ether (PBDE) (Rahman et al., 2001). Concern has been expressed about direct application of the TEF/TEQ approach to abiotic matrices, such as soil, sediment, etc., for human risk assessment. This is problematic as the present TEF scheme and TEQ methodology are primarily intended for estimating exposure and risks via oral ingestion (e.g., by dietary intake).

VII.2.2 Dioxin/furan and PCB TEQ in Bayou Grande Sediments Dioxin/furan analyses do not appear to have been done at NAS or anywhere else in or near Bayou Grande. An important rationale for performing these analyses for this study was to support a related PERCH study of dioxins/furans and dioxin-like PCBs in local seafood. Aroclors of PCBs have been detected on the NAS side of Bayou Grande and have been linked to activities at NAS (EnSafe, 2003; 2004). Other studies have also detected PCB congeners and Aroclors in Bayou Grande (DeBusk et al., 2002).

In the present study, the dioxins/furans contributed on average 59% and the PCBs about 41 % of the total combined TEQ (Table 27). This difference is illustrated by Figure 27 and Figure 28. In the main channel of Bayou Grande the dioxin TEQs tend to be higher than that of the dioxin-like PCBs; the dioxin-like PCB TEQ dominates only in the upper reaches of some embayments. Samples GF-3, 6, 15,18, and 23 had the highest dioxin-like PCB TEQ and were mostly located in or about embayments.

Table 27. TEQ for dioxins/furans, dioxin-like PCBs, and combined total TEQ [ng/kg]. Sample ID Dioxin TEQ PCB TEQ Total TEQ GF-1 2.14 5.8 7.95 GF-2 5.01 1.93 6.94 GF-3 9.82 7.06 16.88 GF-4 6.67 1.83 8.51 GF-5 9.46 0.01 9.48 GF-6 8.54 7.51 16.05 GF-7 10.01 0.01 10.02 GF-8 10.74 0.63 11.37 GF-9 7.78 0.22 7.99 GF-10 0.64 0.12 0.76 GF-11 0.22 0.09 0.31 GF-12 0.09 0.21 0.3 GF-13 8.31 0.21 8.52 GF-14 6.13 1.12 7.25 GF-15 7.62 11.69 19.31 GF-16 0.52 9.22 9.75 GF-17 5.85 0.01 5.86 GF-18 3.05 11.44 14.5 GF-19 0.34 0.63 0.97 GF-20 0.09 0.43 0.52 GF-21 0.74 0.43 1.17 GF-22 1.9 3.41 5.31 GF-23 1.01 10.65 11.66 Mean 4.64 3.25 7.89 1: Boldfaced font indicates that concentration exceeds AET of 3.6 ng/kg.

The TEQ values for the samples ranged from almost non-detect to 10.74 ng/kg for dioxins/furans and 11.69 ng/kg for PCBs. The total combined TEQ ranged from 0.3 to 19.31 ng/kg (Table 27). Seventeen of the 23 surface grabs had combined TEQ values that exceeded the NOAA AET of 3.6 ng/kg. Seven of these samples had a combined TEQs of 10 ng/kg or higher. The three highest combined concentrations were from samples in the small tributary bayous of Bayou Grande (Figure 26). The next four highest concentrations were encountered in or adjacent to the main bayou channel (GF-23, GF-7, GF-8, and GF-18). This suggests that the major embayments of Bayou Grande are the most highly contaminated with TEQ toxicity exceeding SQAGs, but also that the majority of Bayou Grande is generally contaminated. The lowest TEQs included GF-20 in the most western collecting location. Figure 26 shows that the combined TEQ toxicity is not distributed in a homogenous manner in the various regions of Bayou Grande, but the combined TEQ values are generally higher in the eastern half of the Bayou than in the western portion.

Figure 26. Combined TEQ of dioxins/furans and PCBs.

Figure 27. Spatial distribution of TEQ for dioxins/furans.

Figure 28. Spatial distribution of TEQ for dioxin-like PCBs.

A survey for dioxins/furans in Florida Panhandle Bay Systems found TEQs of 23.6 and 13.57 ng/kg in Pensacola Bay and Santa Rosa Sound respectively (Hemming et al., 2003). These values are higher than the TEQs observed by the current study in Bayou Grande. For Eleven Mile Creek in the adjacent Perdido Bay System, that is thought to be impacted by a paper mill, TEQs range from 18.9 to 77.51 ng/kg. In bottom sediments of the Houston Shipping Channel dioxin/furan TEQ concentrations varied from 0.9 to 139.8 ng/kg (Suarez et al., 2006). A comprehensive evaluation of sediment quality in Casco Bay near Portland, Maine detected dioxins/furans in all 30 sediment samples. The median TEQs found for different parts of the bay system were 15 ng/kg in Inner Bay, 13 ng/kg in East Bay, 11 ng/kg in Outer Bay, 5.3 ng/kg in West Bay, and 1.8 ng/kg in Cape Small (Wade et al., 1997). In Bayou Grande dioxin/furan TEQs remain below 11 ng/kg, which shows that compared to other contaminated estuaries dioxin/furan concentrations in Bayou Grande are moderate. Even on a local basis there are considerably higher levels of these compounds in other areas (e.g. Perdido Bay and River System). Another PERCH study, for nearby Bayou Chico, also found higher levels for dioxins/furans (Mohrherr et al., 2006).

VII.2.3 Total Dioxin/Furan Mass Concentrations in Surface Sediments The mean for dioxin/furan mass concentrations in surface sediment was 1,471 ng/kg (Table 28). The highest mass concentration was 3,622.6 ng/kg (GF-8) and occurred in the channel of the Bayou (Figure 29). The highest dioxin/furan concentrations generally occur in the channel in the mid section of the main body of the Bayou as well as in the larger embayments, a distribution that matches that of the dioxin/furan TEQ (Figure 27). This distribution differs from that of the PAHs and PCBs which have the highest concentrations near the sources of embayments (Navy Point Bayou, Yacht Basin, and also Redoubt Bayou for PCBs) (Figure 23, Figure 29).

Dioxins/furans can be formed by many processes including natural forest fires and combustion in human/industrial activities, and thus may reach the Bayou, among other routes, via aerial deposition. Some PCBs are of atmospheric origin, but high sediment concentrations of PCBs are predominantly associated with human land-based activities, as seems to be the case for the PAHs in Bayou Grande. Often there is a discrete point of origin that is associated with releases of PCBs from transformers. These observations suggest that the PCBs and PAHs come into the Bayou via the embayments, and spread from there, while the dioxins/furans may not be exclusively associated with on-shore point sources. There is no evidence to suggest that the wider distribution of the dioxins/furans is due to a greater mobility.

Table 28. Dioxin/furan mass and TEQ concentrations [ng/kg] in surface sediments. Sample Mass TEQ Ratio GF-1 862.5 2.14 2.5 GF-2 1403.6 5 3.6 GF-3 2337.2 9.82 4.2 GF-4 2088.6 6.67 3.2 GF-5 2909.9 9.46 3.3 GF-6 2123.8 8.54 4 GF-7 3196.2 10.01 3.1 GF-8 3622.6 10.74 3 GF-9 2567.3 7.78 3

Table 28. Dioxin/furan mass and TEQ concentrations [ng/kg] in surface sediments - continued.

Sample Mass TEQ Ratio GF-10 216.3 0.64 3 GF-11 101.4 0.22 2.2 GF-12 53.4 0.09 1.7 GF-13 2829.8 8.31 2.9 GF-14 1967.2 6.13 3.1 GF-15 2601.1 7.62 2.9 GF-16 205.4 0.52 2.5 GF-17 1922.5 5.85 3 GF-18 980.4 3.05 3.1 GF-19 109 0.34 3.1 GF-20 55.7 0.09 1.6 GF-21 317.7 0.74 2.3 GF-22 978.9 1.9 1.9 GF-23 390.8 1.01 2.6 mean 1471.4 4.64 3.2

Figure 29. Total dioxin/furan mass concentrations in surface sediments.

VII.2.4 Total Dioxin/Furan Mass Concentrations in Vibracores The highest concentrations for dioxin/furan mass and TEQ from a vibracore sample was from sample GV-1A in Navy Point Bayou (Table 29, see Figure 21 for location). The next four highest mass concentrations were also from the surface (level A), i.e. from GV-11A, 6A, GV-7A

and GV-2A. These concentrations were one to several times lower than what was detected for sample GV-1A. The AET SQAG for dioxin/furan TEQ is surpassed in only two vibracore samples, GV-1A and GV-2B. GV-1 is located near where other high concentrations for surface SOCs were found. Sample GV-11 from the main channel of Bayou Grande had relatively higher mass concentrations down to three meters than were observed for the other vibracores.

Table 29. Dioxin/furan mass and TEQ concentrations [ng/kg] in sediment from vibracores. 0 m1 1 m 2 m 3 m sample mass TEQ mass TEQ mass TEQ mass TEQ GV-1 2682.5 7.21 116.4 0.34 547.4 1.05 GV-2 569.7 1.14 332 5.75 470.1 3.23 GV-4 153.0 0.45 219.8 0.42 15.5 0.02 9.7 0.01 GV-5 78.8 1.06 23.1 0.03 82.0 0.09 GV-6 744.5 1.81 60.5 0.13 206.4 0.31 GV-7 662.9 1.34 32.3 0.04 30.3 0.03 GV-8 85.6 0.19 116.2 0.26 GV-9 123.6 0.3 11.9 0.01 13.6 0.01 9.67 0.02 GV-10 30.0 0.24 12.0 0.15 247.7 0.2 GV-11 1518.3 3.4 539.6 1.1 1439.9 2.8 1167.1 2.9 1: Depth levels correspond to sample A (0 m), B (1 m), C (2 m), and D (3 m).

VII.2.5 Origin of Dioxins/Furans in Bayou Grande There are many potential sources for environmental dioxins/furans. Dioxins/furans are not intentionally produced and are generated as by-products and released from a number of activities. These activities are a broad range of industrial processes that include: • Combustion processes (e.g., municipal, medical and hazardous waste incineration, wood combustion, cement kiln operations, fuel combustion (particularly diesel), steel plant operations and metallurgical processes) • Liquid processes (e.g., pulp bleaching with chlorine or chlorine containing compounds, chlorinated organic chemical manufacturing, and chlorine production) • Chemical manufacture, use and dispersion of contaminated commercial products such as formulated pesticides and chlorinated solvents (e.g., 2,4-D or pentachlorophenol used as a wood preservative) (ATSDR, 1998).

In the present study, profiles of the congeners of the dioxins/furans were constructed to assess their origin. The three major congeners most commonly detected in this study were OCDD, HpCDD and OCDF (Table 30).

Table 30. Average dioxin/furan congener composition for Bayou Grande surface sediments. Dioxin/Furan Congener % Composition Octachlorodibenzo-p-dioxin (OCDD) 86.77 1,2,3,4,6,7,8-Heptachlorodibenzo-p-dioxin (HpCDD) 9.43 Octachlorodibenzofuran (OCDF) 1.33 1,2,3,4,6,7,8-Heptachlorodibenzofuran (HpCDF) 1.09 1,2,3,7,8,9-Hexachlorodibenzo-p-dioxin (HxCDD) 0.47 1,2,3,6,7,8-Hexachlorodibenzo-p-dioxin (HxCDD) 0.29 2,3,4,6,7,8-Hexachlorodibenzofuran (HxCDF) 0.16 1,2,3,6,7,8-Hexachlorodibenzofuran (HxCDF) 0.10 1,2,3,4,7,8-Hexachlorodibenzo-p-dioxin (HxCDD) 0.09 1,2,3,4,7,8-Hexachlorodibenzofuran (HxCDF) 0.09 2,3,4,7,8-Pentachlorodibenzofuran (PeCDF) 0.06 1,2,3,7,8-Pentachlorodibenzo-p-dioxin (PeCDD) 0.04 1,2,3,4,7,8,9-Heptachlorodibenzofuran (HpCDF) 0.04 2,3,7,8-Tetrachlorodibenzofuran (TCDF) 0.02 1,2,3,7,8-Pentachlorodibenzofuran (PeCDF) 0.02 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) 0 1,2,3,7,8,9-Hexachlorodibenzofuran (HxCDF) 0 Total 100

To examine the potential origin of the dioxins/furans in Bayou Grande sediments we used principal component analysis (PCA). In this context, PCA is a statistical method that helps to identify commonalities in the dioxin/furan profiles of various samples, and thus may help identify samples of similar origin. We ran a Q-mode PCA on a dataset that included the congeners of all the sediment samples and of dioxin/furan profiles of known origin. This approach has been used successfully to identify the origin of dioxins/furans in several other studies (e.g., Muller et al., 1999; Birch et al., 2007). The data were transformed logarithmically (log10) and for concentrations below the detection limit half of the detection limit was used. A varimax rotation was applied in the PCA to facilitate interpretation of the components. The dioxin/furan profiles of known origin were obtained from the literature (USEPA, 2006) and were for PCP, forest fires, effluent from wastewater treatment plants, oil fired industrial boilers, barrel burning, landfill flares, combustion of leaded gasoline in cars, combustion of unleaded gasoline, combustion of diesel in cars, combustion of diesel in trucks, coal fired power plants, and municipal waste incinerators. The PCA resulted in five components with an eigenvalue >1. A graphical representation of the PCA results shows some grouping of the sediment samples, and a similarity between the profiles of these groups and some of the profiles of known origin (Figure 30). To focus the PCA, the procedure was run again on all sediment samples and the four profiles of known origin that were most similar to sediment samples in the first PCA run (PCP, forest fires, oil fired industrial boiler, wastewater effluent).

Factor loading plot all profiles of known origin 1

0.8 sediments forest fires PCP known origin 0.6 oil indust. boiler waste inciner.

diesel trucks unleaded 0.4

Component 2 coal power plants ww effluent leaded diesel cars 0.2 landfill flare

barrel burning 0 -0.2 0 0.2 0.4 0.6 0.8 1 Component 1

Figure 30. Factor loading plot for first two principal components for dioxins/furans - all profiles. Two components represent 75 % of total variance in dataset.

The second run of the PCA resulted in four components with an eigenvalue >1. The graphical representation of the results for this second run show distinct groups of sediment samples. One group consists exclusively of GF series grab samples, and has dioxin/furan profiles that are similar to that of PCP (Figure 31). A second group of sediment samples consists mostly of GF series samples and is associated with dioxins/furans profiles originating from forest fires. A third group consists of some GF series samples, vibracore samples from the surface level (A level), and a few vibracore samples from deeper levels. This group is associated with dioxin/furan profiles that derive from oil burning in industrial processes. Two more very distinct groups exist (Figure 31). These groups consist exclusively of samples from deeper sediment levels (B, C, D level) and have dioxin/furan profiles that are similar to those that have been found in wastewater effluent elsewhere (USEPA, 2006). In the existing environmental literature for Bayou Grande we have not found evidence that effluent from wastewater treatment plants has been discharged to the Bayou, although it has been discharged to nearby areas of Pensacola Bay from NAS. Industrial wastewater has been released to the Bayou but it may have very different dioxin/furan profiles. Therefore, it is possible that the statistical similarity between the profiles of the deeper samples and wastewater effluent is due to chance (see below).

Factor loading plot selected profiles of known origin 1

GF18 GF17 group 2 0.9 GF8 GF7 GF2 sediments GF9 GF15 GF5 GF13 GF14 GF6 GF22GF1 GF10 GF3 GF4 known origin GF16 GV1A GF19 group 3 0.8 PCP GF23 forest fires GF21 GV4A GF11 GV9A GV2A group 1 GV11A GV2C GV7A GF20 0.7 GV1C GV6A GV10A oil indust.boiler GV5A GF12 GV8A group 4 GV1B 0.6 GV4B GV11B GV9B Component 2 GV11C 0.5 ww effluent GV9C GV6C GV10B GV4C GV2B GV10C 0.4 GV8B GV7B GV6B GV4D group 5 GV5B GV5C 0.3 GV7C 0.3 0.4 0.5 0.6 0.7 0.8 0.9 1 Component 1

Figure 31. Factor loading plot for first two principal components for dioxins/furans - selected profiles. Two components represent 89.5 % of total variance in dataset.

The GF series samples with dioxin/furan profiles similar to that of PCP are located in the deeper parts of the Bayou and its main embayments. PCP itself was not detected in the sediments and we know of no historical wood treating activities near Bayou Grande that used PCP. It is possible that treated wood with PCP has been employed at NAS or other sites along the Bayou Grande shoreline, or that PCP was used in antifouling paints employed in naval or civilian watercrafts.

The two groups with the deeper sediment samples (GV series, B, C, D level) have dioxin/furan profiles that are different from those of the surface sediments at the respective sites (GV series, A level) (Figure 31). These surface samples from the vibracores, which mostly plot in group 3, integrate sediments from a greater depth range than the surface grab samples (GF series), and generally have dioxin/furan profiles that are intermediate between those of the grab samples and the deeper samples. These systematic differences in profiles associated with sampling depth suggests that stratification of dioxin/furan profiles is present in the sediments. This stratification may be linked to different origins for the dioxins/furans, as indicated by the outcome of the PCA (Figure 31). If this contention is correct, combustion in industrial processes was a major source for the dioxins/furans but more recently forest fires became the dominant source. The association of group 1 with PCP seems to indicate that PCP is currently the prevailing source but it is also possible that group 1 represents spatial differentiation in the origin of the dioxins/furans. A

potential alternative explanation for stratification in the dioxin/furan profiles is a more advanced dechlorination of the dioxins/furans at greater depth, where sediments and their pollutants are older. Advanced dechlorination at depth is supported by a comparison of the relative proportion of the various dioxin/furan homologues from the depth and surface samples of the GV series (Figure 32). The advanced degradation at depth alters the dioxin/furan profiles and, for instance, could by chance make them more similar to profiles observed in wastewater effluent.

80 ] 60

Surface 40 Subsurface

20 Proportion ofconc. mass [%

0 Tetra Penta Hexa Hepta Octa dioxin/furan homologues

Figure 32. Average dioxin/furan homologue profile for GV samples. Homologue proportion is expressed as a proportion of total PCB mass concentration. Values for subsurface samples is mean for all B, C, and D level samples.

VII.2.6 Total PCB Concentrations in Surface Sediments The presence of PCBs in sediments on the NAS side of Bayou Grande has been well documented (EnSafe, 2003; 2004). An interest of the current study was to determine the general distribution of PCB congeners in Bayou Grande and to evaluate if there were any other probable sources for the PCBs other than NAS. Analytical results for the PCBs were compared to FDEP sediment guidelines, which presently list a TEL (21.55 ug/kg) and PEL (188.79 ug/kg) for PCBs based upon the total amount present without regard to the number of congeners analyzed for (McDonald, 1994a, b). In 15 samples out of 23 the TEL was exceeded and one of the samples exceeded the PEL (GF-15, 193.38 ug/kg) (Table 31; Figure 33). This sample was located at the most northern extension of Navy Point Bayou near a stormwater outfall draining surface waters from the Warrington area. Sample GF-14 from the same area was also high (154.08 ug/kg) but sample GF-19 (7.69 ug/kg) from elsewhere in Navy Point Bayou was not. Sample GF-23 at the mouth of this embayment was also high (76.52 ug/kg). Davenport Bayou, which is separate but adjacent to Bayou Grande, also had high total PCB concentrations (Figure 33, Table 31). Its

drainage is also from residential areas in the Warrington community. These findings suggest that some PCBs are originating in the Warrington area, and thus not all PCBs in Bayou Grande originate from NAS. Particle size distribution and organic matter content of the sediments have very low correlation coefficients with total PCB concentrations (< 0.3) and thus do not have a substantial influence on the distribution of the PCBs. Sample GF-3 in Woolsey Bayou near NAS had the second highest total PCB concentration (188.79 ug/kg) and almost exceeded the PEL. Some samples in Redoubt Bayou (GF-6) also had high concentrations (130.95 ug/kg). With the history of PCB releases at NAS (EnSafe, 2003; 2004), the high concentrations detected at Woolsey and Redoubt Bayous are not unexpected. Overall the Bayou is impacted by PCBs as judged by many samples with concentrations exceeding the TEL. The western bayou generally has PCB concentrations below the TEL. At Bayou Grande’s junction with Pensacola Bay the PCB concentration was 12.15 ug/kg. This is equivalent to 50% of the TEL and suggests the possibility of transport of contaminated sediment into Pensacola Bay, although other explanations for the presence of these PCBs are possible. The locations of the sites of maximum contamination suggest that the PCBs may be originating from both sides of Bayou Grande. There is a documented history of releases on the NAS side (EnSafe, 2003; 2004) but the current study is the first to show evidence indicating that PCB contamination also originates from Warrington. Previous data for total PCBs from Bayou Grande (Table 8, Debusk et al., 2002) show the mean total PCB concentration at 44.3 ug/kg with a maximum of 249.5 ug/kg. The PCB results from the current study have an average of 64.1 ug/kg with a maximum of 193.4 ug/kg. These values are within the same range suggesting that total surface PCB concentrations have not undergone major changes in the last 10 to 15 years.

Table 31. Total PCB concentrations [ug/kg] in surface sediments. ID PCBs ID PCBs GF-1 12.15 GF-13 90.32 GF-2 79.84 GF-14 154.08 GF-3 188.22 GF-15 193.38 GF-4 58.68 GF-16 9.11 GF-5 23.58 GF-17 113.94 GF-6 130.95 GF-18 10.06 GF-7 36.84 GF-19 7.69 GF-8 34.53 GF-20 7.31 GF-9 29.54 GF -21 50.18 GF-10 10.54 GF -22 74.11 GF-11 9.46 GF -23 76.52 GF-12 17.42 1: TEL is 21.55 ug/kg and PEL 188.79 ug/kg 2: Boldfaced font indicates concentration exceeds the TEL 3: Boldfaced underline font indicates concentration exceeds the PEL.

Figure 33. Total PCB concentrations in surface sediments.

In a recent study of nearby Choctawhatchee Bay four total PCB sediment concentrations were reported as 221.6, 34.43, 23.6, and 103.8 ug/kg (Hemming et al., 2005). The highest of these concentrations is above the PEL and the other three are above the TEL. This shows that there are other locales in the region that also have high concentrations of PCBs. There was no report of the individual PCB congeners that were analyzed for in this study. In a study of the Houston Shipping Channel concentrations for the 209 congeners ranged from 4.18 to 4600 ug/kg dry wt (Howell et al., 2008). Elsewhere, the EPA Gulf Ecology Division monitored estuaries in the Louisianian Province from 1991 to 1994 to assess ecological conditions on a regional scale (Macauley et al., 1999). It was found that over the four years of monitoring 25.6% of Gulf of Mexico estuarine sediments displayed poor biological conditions, as measured by diverse metrics. This study included twenty-two PCB congeners. Concentrations for total PCBs (the sum of the congeners) ranged from 0 - 299 ug/kg. Less than 1% of the observations exceeded the ER- L guideline of 22.7 ug/kg, which is only slightly higher than the 21.55 FDEP TEL. In a study of San Francisco Bay for selected PCB congeners the following concentrations were obtained: for San Pablo Bay 11.4 ug/kg, Oakland 30.1 ug/kg, and Islais Waterway 164.1 ug/kg. The lowest detection was 5.7 ug/kg for San Pablo Bay and the highest was 255.3 ug/kg for the Islais Waterway (Chapman et al., 1987). In summary, compared to other contaminated estuaries the PCB concentrations found by the present study in Bayou Grande are moderately elevated.

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VII.2.7 Total PCB Concentrations in Vibracores PCBs were generally detected at all vibracore levels (Table 32). SQAGs were exceeded in vibracore GV-4 from Woolsey Bayou near NAS at all sampled depths. This vibracore was located in the general vicinity of the second highest PCB concentration for surface sediments (Table 31). The second highest PCB concentration was measured in a sample from GV-1 from Navy Point Bayou. These observations indicate that these embayments of Bayou Grande are more polluted with PCBs than the Bayou itself.

PCBs were only commercially produced during a limited time in the United States, from 1929 until 1977, and were yet found to be present from the surface to 3 m depth at substantial concentrations in Woolsey Bayou (GV-4) and at detectable concentrations elsewhere in the Bayou. We have not encountered any information on when PCBs were first employed in the Bayou Grande watershed. Lead 210 dating of GV-4, performed with gamma spectrometry at a contracted private lab (Microanalytica), shows that sediments at 3 m depth date to 1930 ± 8.5 years. This suggests that the PCBs at the deeper levels may have originated in the early years of PCB production and remained in the rapidly accumulating sediments, but downward movement from shallower and more recent levels is possible, especially since the PCBs at lower levels tend to be the less chlorinated forms that should be more mobile in the sediments. However, PCBs are hydrophobic and, regardless of chlorination, in general have low aqueous solubility. Consequently, migration from contaminated upper level sediments would likely be low. However, if dissolved macromolecules or colloidal particles are dispersed in the porewater they may sorb the PCBs, allowing them to be transported with these mobile carriers. Chatzikosma and Voudrias (2007) did modeling simulation of PCB transport in terrestrial soils and suggested that PCBs can migrate to greater distances than predicted by using the conventional transport models, due to PCB’s partitioning into the dissolved organic matter (DOM) of water. These studies were for terrestrial soils and consist of modeling and require experimental observations for confirmation. The presence of naphthalenes in deep bayou sediments also presents another possible medium of PCB transport. What appeared to be PCB contaminated transformer oils have been found at waste sites on NAS in the presence of naphthalene and other solvents (EnSafe, 2003; 2004) and these substances may have facilitated PCB transport.

VII.2.8 Origin of PCBs in Bayou Grande Some PCB congeners co-elute during analysis, resulting in only one value for two or more congeners. Co-elute means that during the analysis one or more compounds have identical retention times and the peaks for the compounds run together resulting in just one concentration for two or more congeners. These co-elutions result in a total of 156 separate elutions for the 209 PCB congeners (Table 33). In the present study, the first 33 elutions have relative abundances of more than 0.748 % each and together make up 81 % of the total PCB content. The remaining 123 PCB elutions comprise 19% of the total (Table 33). PCB 187 was the most abundant PCB comprising 11% of the total average for PCBs. This congener makes up 5.4% of Aroclor A1260, which was one of the Aroclors detected at the NAS Bayou Grande shoreline (EnSafe, 2003; 2004). The other Aroclors known to be present at NAS only have trace amounts of congener 187.

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Table 32. Total PCB concentrations [ug/kg] for vibracores. sample 0 m 1 m 2 m 3 m 1000x 1000x 1000x 1000x mass TEQ TEQ/Mass mass TEQ TEQ/Mass mass TEQ TEQ/Mass mass TEQ TEQ/Mass GV-1 34.66 2.1427 61.8 3.36 0.0023 0.7 4.51 0.0015 0.3 GV-2 13.72 0.7307 53.3 2.29 0.001 0.4 2 0.0016 0.8 GV-4 643.581 56.65 88 56.04 0.0029 0.1 64.11 0.0013 0 82.1 0.0398 0.5 GV-5 2.06 0.0028 1.4 0.94 0.0006 0.6 1.2 0.0009 0.8 ------GV-6 6.86 0.0089 1.3 1.2 0.0009 0.8 1.24 0.0008 0.6 ------GV-7 5.97 0.0098 1.6 1.32 0.001 0.8 1.61 0.0008 0.5 ------GV-8 0.0029 1.4 3.44 0.0021 0.6 ------GV-9 0.0276 3.5 1.22 0.0005 0.4 1.85 0.0067 3.6 0.88 0.0004 0.5 GV-10 3.41 0.0036 1.1 1.41 0.0006 0.4 1.97 0.0045 2.3 ------GV-11 30.752 0.0247 0.8 9 0.004 0.4 11.3 0.0051 0.5 7.99 0.0017 0.2

1: Boldfaced underline font indicates concentration exceeds the PEL of 188.79 ug/kg. 2: Boldfaced font indicates concentration exceeds the TEL of 21.55 ug/kg .

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Table 33. Average PCB congener profile for surface sediment in Bayou Grande. IUPAC Cong. [%] IUPAC Cong. [%] IUPAC Cong. [%] IUPAC Cong. [%] IUPAC Cong. [%] 187 11.4 202 0.75 64 0.23 103 0.066 188 0.0066 180,193 7.0 136 0.71 167 0.23 3 0.060 54 0.0055 153,168 5.6 85,116,117 0.70 17 0.21 27,24 0.060 111 0.0049 129,138,160, 5.4 20,28 0.70 130 0.21 2 0.060 80 0.0048 163 134,147,149 4.7 128,166 0.67 8 0.21 159 0.058 155 0.0046 194 4.1 171,173 0.62 56 0.21 137 0.053 63 0.0045 198,199 3.9 44,47,65 0.58 209 0.20 107,124 0.051 35 0.0045 203 3.0 183 0.57 22 0.17 126 0.045 36 0.0034 110,115 2.9 31 0.56 205 0.15 169 0.042 78 0.0031 170 2.4 190 0.52 82 0.15 59,62, 75 0.038 127 0.0021 206 2.4 66 0.50 32 0.15 19 0.036 9 0.0016 90,101,113 2.3 156,157 0.49 144 0.15 67 0.031 181 0.0013 135,151,154 2.1 201 0.47 42 0.14 120 0.028 12,13 0.0011 118 1.9 49,69 0.47 26,29 0.14 114 0.024 121 0.0009 196 1.8 164 0.45 109 0.13 152,150 0.021 57 0.0006 177 1.7 208 0.45 98,102 0.13 5 0.020 143 0.0003 11 1.6 172 0.44 16 0.13 131,142 0.019 55 0.0002 95,93,100 1.5 21,33 0.42 45,51 0.13 94 0.019 165 0.0001 132 1.4 197,200 0.40 77 0.12 79 0.017 145 0.0001 178 1.4 88,91 0.40 4 0.10 6 0.015 10 0.0000 174 1.3 41,71,40 0.36 50,53 0.10 122 0.015 14 0.0000 195 1.2 106 0.35 133 0.09 72 0.014 34 0.0000 175 1.2 158 0.32 1 0.09 68 0.014 23 0.0000 83,99 1.2 92 0.30 189 0.09 89 0.013 39 0.0000 179 1.2 207 0.30 25 0.09 46 0.012 38 0.0000 86,87,97,108 1.2 15 0.29 191 0.09 58 0.011 123 0.0000 185 1.2 84 0.29 204 0.08 81 0.010 161 0.0000 70,61,74,76 1.0 37 0.28 48 0.08 7 0.009 184 0.0000 141 0.8 176 0.25 60 0.08 148 0.009 186 0.0000 52,43,73 0.8 112 0.25 139,140 0.07 96 0.009 182 0.0000 105 0.8 18,30 0.24 162 0.07 104 0.007 192 0.0000 146 0.75 1: Boldface indicates PCBs with dioxin-like activity.

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To examine the potential origin of the PCBs in Bayou Grande we ran a Q-mode PCA, similar to the one used for dioxins/furans. Results did not show any systematic grouping of samples that could point to the prevalence of certain profiles or spatial differences (Figure 34). However, the subsurface samples from vibracore GV-4, that had higher total PCB concentrations than other subsurface samples, have congener profiles that are different from those in all other samples. A hierarchical cluster analysis on standardized congener data for surface sediments revealed one cluster in the deeper part of the main bayou, one cluster in the larger embayments in the eastern half of the bayou, and one cluster along the northern shoreline and in the upper reaches of the northern embayments (Figure 35). To further examine the potential origin of the PCBs, and try to explain this spatial pattern, we performed visual comparisons of the PCB homologues for sediments with those in Aroclors.

Factor loading plot

GV4D GV4C GV4B 0.8

GV5B GV2B GV2C 0.6 GV1B GV6C GV5C GV7C GV5A GV9D GV8B GF20 GV6B GV7B GV1C GV4A GF9 GV7A 0.4 GV8A GF11 GV10B GV6AGF1 GF13 GF10 Component 2 Component GF12GV2A GV9B GV11D GV1A GF21 GF22 GF23 GV11B GF16 GV9A GF18GF5GF7 GV10C GF8 0.2 GV9C GV10A GF19 GV11C GF17 GF2 GV11A GF14 GF4 GF15 GF6

0 GF3 0 0.2 0.4 0.6 0.8 1 Component 1

Figure 34. Factor loading plot for first two principal components for PCBs.

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Figure 35. Cluster analysis of PCB congener data for surface sediments.

Results for the surface grab samples (GF series) shows that the homologue patterns for the sediments are most similar to that of Aroclor 1260 (Figure 36a). Even though the match is not perfect, Aroclor 1260 is the only Aroclor known to be present in the Bayou that has a large proportion of hexa- and hepta-CBs, as the sediments do. The imperfect match may be due to degradation of the PCBs or mixing with PCBs from secondary sources. Samples GF 14, 15, 16, 17, and 18 have a larger proportion of penta-CBs and their profile seems to be equivalent to that of a mix of Aroclors 1260 and 1254 (Figure 36a). The surface samples of the GV series follow this same trend: GV-2A, 4A, and 11A have homolog profiles similar to that of Aroclor 1260 and the others have a profile that seems to be a mix of Aroclors 1260 and 1254 (Figure 36b). The samples of both series that seem to represent a mix of Aroclors were grouped in cluster 3 by the hierarchical cluster analysis and are located along the northern shore of the Bayou and its embayments (Figure 35). Aroclor 1260 has been detected at NAS (EnSafe, 2003; 2004) and Aroclor 1254 has been used in more applications than any of the other Aroclors and thus has many potential sources. These observations indicate that the PCBs in much of Bayou Grande may derive from NAS but that along the northern shore, away from NAS and near Warrington and other residential areas, PCBs originate from more than one source. This finding is consistent with results presented in section VII.2.6 that show that PCBs are entering the Bayou from the north and south. Obviously, it cannot be excluded that the Aroclor 1260 like PCB profiles in much of the Bayou originated from other, non-NAS, sources unknown to us.

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PCB homologues for arochlors and GF samples

A1221 A1232 A1016 A1242 A1248 A1254 A1260 GF-23 Tot MoCB GF-22 Tot DiCB GF-21 Tot TriCB GF-20 Tot TeCB GF-19 Tot PeCB GF-18 Tot HxCB GF-17 Tot HpCB GF-16 GF-15 Tot OcCB

Samples GF-14 Tot NoCB GF-13 Tot DeCB GF-12 GF-11 GF-10 GF-9 GF-8 GF-7 GF-6 GF-5 GF-4 GF-3 GF-2 GF-1

0% 20% 40% 60% 80% 100% Proportional composition

Figure 36a. PCB homologue profiles for GF series samples and Aroclors.

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PCB homologues for arochlors and GV-A samples

A1221

A1232

A1016

A1242

A1248

A1254 Tot MoCB A1260 Tot DiCB Tot TriCB GV-11A s Tot TeCB Tot PeCB GV-10A Tot HxCB

Sample GV-9A Tot HpCB Tot OcCB GV-8A Tot NoCB Tot DeCB GV-7A

GV-6A

GV-5A

GV-4A

GV-2A

GV-1A

0% 20% 40% 60% 80% 100% Proportional composition

Figure 36b. PCB homologue profiles for surface sediments from GV series and Aroclors.

The subsurface samples from the GV series do not show a clear similarity with any of the Aroclors with the exception of the GV-4 samples (Figure 36c), which also have congener profiles that are different from those of other samples (Figure 34). The homolog pattern of the GV-4 samples is very similar to that of Aroclor 1221 and has a high proportion of mono- and di-CBs. Total PCB concentrations in these samples are very high (Table 32). These observations indicate that either a large or long term release of PCBs, possibly a spill of an Aroclor 1221 containing substance, may have occurred in or near the headwaters of Woolsey Bayou, where

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GV-4 is located. An argument against the interpretation of Aroclor 1221 being the source for the deeper samples is the presence of some high chlorinated homologues in the samples. These more fully chlorinated homologues are more consistent with natural degradation of more completely chlorinated Aroclors than with an origin from Aroclor 1221. In any case, the most important finding in this context are the elevated concentrations of PCBs in this embayment, even at 3 m depth, regardless of the precise origin on the PCBs.

PCB homologues for arochlors and subsurface sediments

A1260 A1254 A1248 A1242 A1016 A1232 A1221 G V - 11D G V - 11C Tot MoCB G V - 11B GV- 10C Tot DiCB GV- 10B Tot TriCB GV- 9D s GV- 9C Tot TeCB GV- 9B Tot PeCB GV- 8B

Sample Tot HxCB GV- 7C GV- 7B Tot HpCB GV- 6C Tot OcCB GV- 6B Tot NoCB GV- 5C GV- 5B Tot DeCB GV- 4D GV- 4C GV- 4B GV- 2C GV- 2B GV- 1C GV- 1B

0% 20% 40% 60% 80% 100% Proportional composition

Figure 36c. PCB homologue profiles for subsurface sediments from GV series and Aroclors.

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VII.2.9 Dioxin-like PCB Concentrations in Surface Sediments The most toxic effects of PCBs are currently thought to reside with the twelve specific PCB congeners that have dioxin-like effects (Van den Berg et al., 2006). The TEQ for these dioxin- like PCBs relative to the total PCB mass concentration varies considerably between collecting sites in Bayou Grande (Table 34). Congeners 126 and 169 constitute less than 3% of the total dioxin-like PCB mass but are responsible for approximately 76% and 22% respectively of the total TEQ due to their higher TEFs (Table 35, Table 36). Congener 118 comprises 51% of the total mass of the dioxin-like PCBs (Table 37) and about 1% of the total TEQ (Table 36, Table 37). This congener is the only dioxin-like PCB that is a major constituent of an Aroclor and is present at concentrations of 7.35 to 13.5% in Aroclor 1254 (Table 38), which has not been detected in Bayou Grande sediments (EnSafe, 2003; 2004). The value of these findings is that they will be useful in related ongoing PERCH studies for bioaccumulation in seafood and other impacts upon the local biota. The concentrations found in Bayou Grande are similar to those reported for Bayou Chico where PCB 126 and 169 represented 65% and 15% respectively of the TEQs and PCB 118 represented about 53% of the mass concentration of all dioxin-like PCBs (Mohrherr et al., 2006).

Table 34. Dioxin-like PCB TEQ [ng/kg] and total PCB mass concentration [ug/ kg] in surface sediments. Total PCB TEQ/Total PCB Sample TEQ ug mass /kg ug mass /kg3 GF-1 5.8 12.15 477 GF-2 1.93 79.8431 24 GF-3 7.06 188.22 38 GF-4 1.83 58.68 31 GF-5 0.01 23.58 0.4 GF-6 7.51 130.95 57 GF-7 0.0086 36.84 0.2 GF-8 0.63 34.53 18 GF-9 0.22 29.54 7 GF-10 0.11 10.54 10 GF-11 0.09 9.46 10 GF-12 0.21 17.42 12 GF-13 1.12 90.32 12 GF-14 11.69 154.08 76 GF-15 9.22 193.3842 48 GF-16 0.0073 9.11 0.8 GF-17 11.44 113.94 100 GF-18 0.02 10.06 2 GF-19 0.63 7.7 82 GF-20 0.43 7.31 59 GF-21 3.41 50.18 68

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Table 34. Dioxin-like PCB TEQ [ng/kg] and total PCB mass concentration [ug/ kg] in surface sediments - continued.

Total PCB TEQ/Total PCB Sample TEQ ug mass /kg ug mass /kg3 GF-22 10.65 74.11 144 GF-23 9.24 76.52 121 Mean 3.62 61.67 59

1: Bold faced font indicates concentration exceeds the TEL of 21.55 ug/kg. 2: Italicized underline font indicates concentration exceeds the PEL of 188.79 ug/kg. 3 Ratio of sample TEQ ng/kg divided by total sample PCB mass ug/kg x 1000.

Previous studies have found that on the NAS side of Bayou Grande Aroclor 1260 has more detections than other Aroclors (82 vs. 5) and that the mean concentrations of Aroclor 1260 are higher (EnSafe, 2003; 2004) (Table 7). Congener 187 is present in Aroclor 1260 and is the most common non-dioxin like congener in Bayou Grande. This suggests that Aroclor 1260 may have been the source of the most common non-dioxin-like PCB congener in the Bayou. This contention is consistent with the findings for the origin of the PCBs based on cluster analysis and a comparison of homologue profiles (see above). Aroclor 1254, however, may have contributed the most common dioxin-like congeners such as congener 118 and 105 which are low in Aroclor 1260.

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Table 35. Dioxin-like PCB congener mass concentration [ng/kg] for surface sediments. Sample 771 81 123 118 114 105 126 167 156&157 169 189 Sum GF-1 47 46 ND 140 51 68 45 53 104 42 52 648 GF-2 49 ND ND 1740 17 619 18 207 469 ND 96 3215 GF-3 58 15 ND 1090 18 250 56 415 711 44 251 2908 GF-4 48 ND ND 487 5 157 12 75 138 20 30 974 GF-5 12 ND ND 185 ND 49 ND 26 48 ND 9 329 GF-6 189 ND ND 2108 19 662 69 294 549 15 140 4044 GF-7 12 ND ND 153 ND 37 ND 14 34 ND 10 259 GF-8 31 ND ND 342 4 85 6 34 64 ND 13 578 GF-9 16 ND ND 214 2 58 ND 16 31 7 5 349 GF-10 3 ND ND 42 ND 13 ND 4 8 4 2 75 GF-11 2 ND ND 36 ND 9 ND 4 6 3 2 62 GF-12 1 ND ND 42 ND 12 ND 4 7 7 ND 73 GF-13 26 ND ND 498 ND 128 ND 46 95 36 21 850 GF-14 261 ND ND 3284 62 1424 104 435 910 36 145 6661 GF-15 360 61 ND 7958 132 3893 77 707 1958 33 162 15344 GF-16 13 ND ND 114 ND 45 ND 9 23 ND 7 212 GF-17 283 ND ND 4463 ND 1874 112 397 885 ND 97 8110 GF-18 27 ND ND 259 4 91 ND 25 48 ND 7 460 GF-19 14 ND ND 208 ND 84 6 21 45 0 8 387 GF-20 6 ND ND 50 ND 14 ND 6 12 14 ND 102 GF-21 47 ND ND 429 ND 103 ND 47 94 113 22 855 GF -22 141 ND ND 1473 ND 459 73 214 396 107 99 2962 GF -23 122 ND ND 1485 21 517 58 170 370 111 71 2926 Mean 77 5 0 1165 15 463 28 140 305 26 54 2278 % of tot. 3.38 0.23 0 51.16 0.64 20.33 1.22 6.15 13.37 1.13 2.39 100 PCBs

1: IUPAC number.

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Table 36. Dioxin-like PCB congener TEQs [ng/kg] for surface sediments sample 771 81 123 118 114 105 126 167 156&157 169 189 Total GF-1 0.005 0.014 0.000 0.004 0.002 0.002 4.510 0.002 0.003 1.262 0.002 5.80 GF-2 0.005 0.000 0.000 0.052 0.001 0.019 1.835 0.006 0.014 0.000 0.003 1.93 GF-3 0.006 0.004 0.000 0.033 0.001 0.008 5.642 0.012 0.021 1.328 0.008 7.06 GF-4 0.005 0.000 0.000 0.015 0.000 0.005 1.193 0.002 0.004 0.608 0.001 1.83 GF-5 0.001 0.000 0.000 0.006 0.000 0.002 0.000 0.001 0.001 0.000 0.000 0.01 GF-6 0.019 0.007 0.000 0.063 0.001 0.020 6.924 0.009 0.017 0.447 0.004 7.51 GF-7 0.001 0.000 0.000 0.005 0.000 0.001 0.000 0.000 0.001 0.000 0.000 0.01 GF-8 0.003 0.000 0.000 0.010 0.000 0.003 0.614 0.001 0.002 0.000 0.000 0.63 GF-9 0.002 0.000 0.000 0.006 0.000 0.002 0.000 0.001 0.001 0.205 0.000 0.22 GF-10 0.000 0.000 0.000 0.001 0.000 0.000 0.000 0.000 0.000 0.116 0.000 0.12 GF-11 0.000 0.000 0.000 0.001 0.000 0.000 0.000 0.000 0.000 0.090 0.000 0.09 GF-12 0.000 0.000 0.000 0.001 0.000 0.000 0.000 0.000 0.000 0.209 0.000 0.21 GF-13 0.003 0.000 0.000 0.015 0.000 0.004 0.000 0.001 0.003 1.093 0.001 1.12 GF-14 0.026 0.000 0.000 0.099 0.002 0.043 10.389 0.013 0.027 1.088 0.004 11.69 GF-15 0.036 0.018 0.000 0.239 0.004 0.117 7.735 0.021 0.059 0.990 0.005 9.22 GF-16 0.001 0.000 0.000 0.003 0.000 0.001 0.000 0.000 0.001 0.000 0.000 0.01 GF-17 0.028 0.000 0.000 0.134 0.000 0.056 11.183 0.012 0.027 0.000 0.003 11.44 GF-18 0.003 0.000 0.000 0.008 0.000 0.003 0.000 0.001 0.001 0.000 0.000 0.02 GF-19 0.001 0.000 0.000 0.006 0.000 0.003 0.613 0.001 0.001 0.000 0.000 0.63 GF-20 0.001 0.000 0.000 0.002 0.000 0.000 0.000 0.000 0.000 0.429 0.000 0.43 GF-21 0.005 0.000 0.000 0.013 0.000 0.003 0.000 0.001 0.003 3.381 0.001 3.41 GF-22 0.014 0.000 0.000 0.044 0.000 0.014 7.344 0.006 0.012 3.215 0.003 10.65 GF-23 0.012 0.000 0.000 0.045 0.001 0.016 5.811 0.005 0.011 3.338 0.002 9.24 Mean 0.008 0.002 0.000 0.035 0.001 0.014 2.774 0.004 0.009 0.774 0.002 3.621 % of tot. 0.21 0.05 0.00 0.97% 0.01 0.39 76.60 0.11 0.25 21.37 0.04 100.00 PCBs

1: IUPAC number.

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Table 37. Average dioxin-like PCB mass concentration and TEQ. Cong.TEQ/ Cong. mass/ Cong. mass/ IUPAC Formula TEF Tot. diox-like Tot.diox-like Total PCB No PCB mass PCB mass mass 126 3,3',4,4',5-PentaPCB 0.1 76.43% 1.21% 0.04% 169 3,3',4,4',5,5'-HexaPCB 0.03 21.53% 1.14% 0.04% 118 2,3',4,4',5-PentaPCB 0.00003 0.96% 51.11% 1.88% 105 2,3,3',4,4'-PentaPCB 0.00003 0.38% 20.31% 0.75% 2,3,3',4,4',5-& 2,3,3',4,4',5' 156&157 0.00003 0.25% 13.38% 0.49% HexaPCB 77 3,3',4,4'-TetraPCB 0.0001 0.21% 3.37% 0.12% 167 2,3',4,4',5,5'-HexaPCB 0.00003 0.12% 6.15% 0.23% 81 3,4,4',5-TetraPCB 0.0003 0.05% 0.28% 0.01% 189 2,3,3',4,4',5,5'-HeptaPCB 0.00003 0.05% 2.39% 0.09% 114 2,3,4,4',5-PentaPCB 0.00003 0.01% 0.65% 0.02% 123 2',3,4,4',5-PentaPCB 0.00003 0.00% 0.00% >0.0001%

Table 38. Dioxin-like PCB congener distributions [%] for seven Aroclor mixtures (modified from Frame et al., 1996). IUPAC 1 1 1 1 A1016 A1242 A1248 A1248 A1254 A1254 A1260 No. 126 0 0 trace trace 0.02 trace 0 169 0 0 0 0 0 0 0 118 0 0.66 2.29 2.35 13.59 7.35 0.48 105 trace 0.47 1.6 1.45 7.37 2.99 0.22 156 0 0.01 0.06 0.04 1.13 0.82 0.52 157 0 0 0.01 trace 0.3 0.19 0.02 77 0 0.31 0.41 0.52 0.2 0.03 0 167 0 0 0.01 0.01 0.35 0.27 0.19 81 0 0.01 0.01 0.02 trace 0 0 189 0 0 0 0 0.01 0.01 0.1 114 0 0.04 0.12 0.12 0.5 0.18 0 123 0 0.03 0.07 0.08 0.32 0.15 0

1: The two columns for Aroclors 1248 and 1254 refer to different formulations for the Aroclor.

VII.2.10 Degradation of PAHs, PCBs, and Dioxins/ Furans in Sediments Concentrations for most pollutants decrease with depth in Bayou Grande sediments. Among the POPs, dioxins/furans declined the least with increasing depth in vibracore samples even though their congener profiles change (Table 21), PAHs declined the most (Table 29). PCB toxicity expressed as the ratio of dioxin-like PCB TEQ per Total PCB mass also decreases considerably with depth (Table 32).

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Degradation of POPs occurs at rates that are influenced by the chemical structure and solubility of the molecules involved and the environmental conditions within the sediments. Degradation of a pollutant molecule will effect what is detected by analytical methods with the degree of the effect being dependent upon the specific analyses. Degradation of PAHs will have a different effect on analytical detections when compared to degradations of PCBs and dioxins/furans. The 8270 C analysis used in the current study detects 18 specific PAH congeners out of a much larger number possible PAH species. After alteration of the parent configuration by changes such as methylation to these 18 molecules may result in a non-detect for the degraded PAH. Only specific chromatogram peaks register as a detection of a PAH congener and while a degraded molecule may still be a PAH, its detection peak is altered and may not register as a detection once its structure and/or mass has been altered. Therefore, while these altered PAHs are not detected they still exist and their impact upon the environment is not considered. The lower concentrations then are likely due to both degradation because of longer residence times in the sediments and also to lower initial concentrations of these PAHs that are associated with anthropogenic activities.

In Table 39 selected representative sediment core PCB, PCB TEQ, and PAH concentrations are compared. For PCBs the typical mass concentration decreases with depth to about 10% of that at the surface (A level), but TEQs decline considerably more. Detectable total PAH concentrations decline by a factor of about 100 or more. Thus in most cases it appears that the likely region of maximum impact from SOCs is upon the surface sediments, a finding that is relevant to issues of sediment remediation and dredging.

Table 39. Selected vibracore PCB, PCB TEQ, and PAH concentrations. Analyte PCB PCB TEQ PAH

GV-4 Ratio Level GV-4 Ratio Level GV-1 Ratio Level Level [ug/kg] to Level A [ng/kg] to Level A [ug/kg] to Level A

A 643.58 1.0 56.65 1.0 1426 1.0 B 56.04 0.09 0.003 0.0001 2.4 0.002 C 64.11 0.10 0.001 0.00002 15.1 0.01 D 82.1 0.13 0.04 0.0007 NS1 NS 1: NS indicates no sample was taken at this level.

Degradation of PCBs and dioxins/furans can occur under anaerobic conditions in deeper sediments where anaerobic conditions favor dechlorination. Dechlorination results in a “degraded” molecule with altered toxicity. The degraded molecule can have a more potent impact than the parent molecule in the case of dioxins/furans. Unlike PAHs detected by EPA Method 8270 C, EPA methods 1613B and 1668A for PCBs and dioxins/furans can detect partially degraded congeners resulting from dechlorination. EPA Method 1668A analyzes all congeners of PCB and regardless of the position of any dechlorination the resulting degradation will be detected as a PCB if any chlorine remains bonded to the biphenyl molecule.

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The possibility of dechlorination of PCBs in Bayou Grande is suggested by the decrease in the TEQ/mass ratio of the dioxin-like PCBs for the deeper levels versus the upper levels (Table 39). Since these ratios are different, the congener profiles of the dioxin-like PCBs most likely also differ. Figure 37, Figure 38, and Figure 39 show the relative concentrations for the totals of the ten PCB homologues (mono, di, tri, tetra, penta, hexa, hepta, octa, nona, and decachlorinated PCBs) in those vibracores with more than trace amounts of PCBs in the lower levels. In all three samples the di, tri, and tetrachlorinated homologues were lowest at level A and higher in the B, C, and D levels. These data show that the decrease with depth in TEQ relative to total PCB mass (Table 39) represents a change in overall chlorination that is likely due to degradation or selective transport.

Vibracore GV1 PCB Homologues At Different Depths

30% Level A Level B 25% Level C

20%

15%

10% Percent Homologue Percent

0% MoCB DiCB TriCB TeCB PeCB HxCB HPCB OcCB NoCB DeCB PCB Homologues

Figure 37. PCB homologues in vibracore GV-1.

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Vibracore GV4 PCB Homologues At Different Depths

60% LeveLA Level B 50% Level C Level D 40%

30%

20% Percent Homologue Percent

10%

0% MoCB DiCB TriCB TeCB PeCB HxCB HPCB OcCB NoCB DeCB PCB Homologues

Figure 38. PCB homologues in vibracore GV-4.

Vibracore GV-11 PCB Homologues At Different Depths

35% Level A Level B 30% Level C Level D 25%

20%

15%

Percent Homologue Percent 10%

0% MoCB DiCB TriCB TeCB PeCB HxCB HPCB OcCB NoCB DeCB PCB Homologues

Figure 39. PCB homologues in vibracore GV-11.

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Biodegradation of PCB congeners has been commonly observed to occur in sediments (Sokol et al., 1998). Biodegradation of PCB congeners occurs via dechlorination by aerobic and anaerobic bacteria. The lighter congeners such as the mono, di and tri PCBs generally biodegrade aerobically. The heavier congeners, e.g., tetra and penta PCBs generally biodegrade anaerobically. Studies have shown that anaerobic processes can reduce total chlorine of PCBs by 33 - 36% over several decades (Sokol et al., 1998). Anaerobic dechlorination depends on meta and para removal of chlorine (Figure 40). The removal of para chlorines occurs with the removal of adjacent chlorine (tetra to hepta PCBs). The removal of meta chlorines occurs with at least one adjacent chlorine for nearly all hexa and hepta PCB congeners resulting in production of lighter molecular weight congeners. The lighter molecular weight PCB congeners are subject to aerobic microbial degradation and can oxidatively mineralize to carbon dioxide and water that since it is aerobic is more likely to occur in surface sediments (Sokol et al., 1998).

Figure 40. Ortho, meta, and para positions on aromatic molecules.

The structure of dioxin-like-PCBs (Figure 9) consists of substituent chlorines occupying: (a) usually no more than one of the ortho positions; (b) both para positions; and (c) at least two meta positions. Polychlorinated biphenyls, including the dioxin-like PCBs, are extremely resistant to conventional aerobic transformation, but they will undergo anaerobic reductive dechlorination and this could conceivably occur in the lower levels of the sediments. For PCBs, the degradation rate is inversely related to the degree of chlorination; thus, highly chlorinated congeners are more readily dechlorinated than congeners with a lower degree of chlorination. Studies of PCB contamination in Hudson River sediment demonstrate that anaerobic environments yield markedly lower levels of tri-, tetra-, and pentachlorobiphenyls and higher levels of mono- and dichlorobiphenyls (Bedard and Quensen, 1995). Many of the dioxin-like PCBs are tetra and pentachlorobiphenyls. Tetra, penta, hexa and hepta PCBs are subject to anaerobic dechlorination in sediments in either the meta or para positions. All dioxin-like PCBs must retain chlorine at the para or “4” positions to retain dioxin-like activity. Reductive dechlorination predominantly reduces chlorine in meta- and para-positions, resulting in accumulation of the ortho-chlorinated congeners. (Fiedler et al., 1994). Additionally, it is well known that light molecular weight chlorinated PCBs can volatilize (Mackay et al., 1992) and this may be another mechanism that

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effects the concentrations of total PCB mass in Bayou Grande and would be expected to effect concentrations at the surface more than at deeper levels.

The dioxin/furan concentrations were generally higher at the surface than at depth, but the contrast was not as marked as for PCBs (Table 29). In samples from vibracore GV-2, for instance, the lowest level at 2 m has a concentration that is not much less than that at the surface (470 ng/kg vs. surface 570 ng/kg). For vibracore GV-11, taken in the Bayou’s basin, level A is 1,518 ng mass/kg, B is 539 ng mass/kg, C is 1,439 ng mass/kg, and D is 1,167 ng mass/kg.

Dioxins/furans unlike PCBs do not exclusively derive from a specific manufacturing activity and can be produced by a variety of circumstances, especially those involving combustion or heat in the presence of chlorine. They have been produced for many years as byproducts of combustion and other processes involving high temperatures without a specific time line, in contrast to the case for PCBs. The dechlorination of dioxins/furans, unlike the PCBs, could likely increase the toxicity resulting in an increased TEQ (Table 41, Figure 41). The most common forms of toxic dioxins/furans are the octachlorinated compounds such as OCDD (octachlorodibenzo-p-dioxin). Removal of chlorine could result in an increase in TEQ due to the fact that as the number of chlorines are removed from the more common OCDD and OCDF (Octachlorodibenzofuran) the TEF for the congeners can increases dramatically depending on which specific congeners are formed (e.g. from a TEF of 1.0 for 1,2,3,7,8-Pentachlorodibenzo-p-dioxin to a TEF of 0.0003 for 2,3,7,8-Tetrachlorodibenzo-p-dioxin) (Table 41, Figure 41). In contrast to PCBs for GV-1, 2, 4 and 11 (Table 32), the mean toxicity on a basis of TEQ/mass for dioxins/furans shows less difference between different sediment depths in Bayou Grande (Table 40).

Table 40. TEQ-to-mass ratio for dioxins/furans in vibracore samples. sample 0 m 1 m 2 m 3 m GV-1 0.271 0.29 0.19 ns2 GV-2 0.20 1.73 0.69 ns GV-4 0.29 0.19 0.13 0.10 GV-5 1.35 0.13 0.11 ns GV-6 0.24 0.21 0.15 ns GV-7 0.20 0.12 0.10 ns GV-8 0.22 0.22 ns ns GV-9 0.24 0.08 0.07 0.21 GV-10 0.80 1.25 0.08 ns GV-11 0.22 0.20 0.19 0.25 Mean 0.40 0.44 0.19 0.19 Mean 1.00 1.10 0.47 0.46 Change 1: (TEQ/Mass)*100 2: ns: not sampled

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Table 41. TEFs for dioxin/furan congeners. TEF Congener 0.0003 Octachlorodibenzo-p-dioxin 0.0003 Octachlorodibenzofuran 0.01 1,2,3,4,6,7,8-Heptachlorodibenzo-p-dioxin 0.01 1,2,3,4,6,7,8-Heptachlorodibenzofuran 0.01 1,2,3,4,7,8,9-Heptachlorodibenzofuran 0.03 1,2,3,7,8-Pentachlorodibenzofuran 0.1 1,2,3,4,7,8-Hexachlorodibenzo-p-dioxin 0.1 1,2,3,6,7,8-Hexachlorodibenzo-p-dioxin 0.1 1,2,3,7,8,9-Hexachlorodibenzo-p-dioxin 0.1 2,3,7,8-Tetrachlorodibenzofuran 0.1 1,2,3,4,7,8-Hexachlorodibenzofuran 0.1 1,2,3,6,7,8-Hexachlorodibenzofuran 0.1 1,2,3,7,8,9-Hexachlorodibenzofuran 0.1 2,3,4,6,7,8-Hexachlorodibenzofuran 0.3 2,3,4,7,8-Pentachlorodibenzofuran 1 2,3,7,8-Tetrachlorodibenzo-p-dioxin 1 1,2,3,7,8-Pentachlorodibenzo-p-dioxin

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Figure 41. Dechlorination of dioxin/furan congeners (from Barkovskii et al., 1996). Shown are several possible pathways for the dechlorination of dioxins/furans that form the more toxic penta and tetra chlorinated forms.

VII.2.11 Dioxins/Furans in Bayou Grande Sediments and Seafood Tissues Concentrations of dioxins/furans have been determined in crabs and oysters in the Pensacola Bay System, including Bayou Grande, by a previous PERCH task (Karouna-Renier et al., 2007). In Bayou Grande oysters were obtained from the Navy Blvd. Bridge and crabs from the upper Bayou (where the Bayou narrows between Weller and Paulding Avenues) and lower bayou (slightly upstream of the Navy Blvd. Bridge). Oyster soft non-muscular tissue and crab muscle and hepatopancreas were analyzed separately (Table 42, Figure 42, Figure 43, Figure 44). Dioxin/furan concentrations were highest in upper bayou hepatopancreas samples. Crab muscle tissue was considerably lower at the same site. Relative to sediment concentrations the seafood tissues do not appear to bioconcentrate dioxins/furans to the same extent as was observed for dioxin-like PCBs (see below). Similar observations were reported in a previous study of nearby Bayou Chico (Mohrherr et al., 2006).

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Table 42. Dioxins/Furans in seafood tissues1 [ng/kg wet wt] and sediments [ng/kg dry weight]. Upper Lower Upper Lower Bayou bayou crab bayou crab Congener bayou crab bayou crab Oyster Grande hepato hepato muscle muscle sediment pancreas pancreas 2378-TCDD 0.3 0.3 0.03 0.02 0.01 0.00 12378-PeD 1.2 1.2 0.06 0.03 0.24 0.64 123478-Hx 0.6 0.6 0.02 0.02 0.25 1.37 123678-Hx 1.5 1.4 0.02 0.02 0.63 4.21 123789-Hx 1 0.8 0.06 0.02 0.54 6.91 1234678-Hp 3 2.4 0.24 0.08 1.78 138.70 OCDD 6 5.2 1.06 0.23 15.1 1276.68 2378-TCDF 3.7 4.1 0.23 0.11 1.08 0.34 12378-PeF 5.2 0.7 0.19 0.01 0.11 0.33 23478-PeF 2.2 2.3 0.11 0.08 0.59 0.84 123478-HxF 1.1 0.3 0.02 0.03 0.01 1.35 123678-HxF 22.2 0.3 0.41 0.01 0.01 1.42 123789-HxF 0.2 0 0.03 0.01 0.02 0.00 234678-HxF 0.1 0.3 0.02 0.01 0.01 2.35 1234678-HpF 55.1 0.4 1.11 0.01 0.14 16.02 1234789-HpF 0.4 0 0.04 0.01 0.02 0.60 OCDF 0.2 0.1 0.31 0.03 0.34 19.62 Sum 104 20.4 3.96 0.73 20.88 1471.37

1: Seafood data are from Karouna-Renier et al. (2007), a publication based on another PERCH task.

Profiles of the dioxin/furan congeners for crab tissues differ considerably from those for oyster tissues (Figure 42). OCDD (octachlorodibenzo-p-dioxin) is the predominant dioxin/furan in oyster tissues, but HpCDF (1,2,3,4,6,7,8-heptachlorodibenzofuran) and HxCDF (1,2,3,6,7,8- hexachlorodibenzofuran) are the predominant congeners in upper bayou crab hepatopancreas.

Dioxins/Furans in Bayou Grande Seafood Tissues 15 Upper Crabs Hepato Lower Crabs Hepato 55 ng HpCDF Upper CrabMuscle Lower Crab Muscle 10 16 ng OCDD Oyster 22 ng HxCDF

ng mass/kg wt ng mass/kg 5

0 TCDF TCDD OCDF OCDD HxCDF HxCDF HxCDF HxCDF HxCDD HxCDD HxCDD PeCDF PeCDF HpCDF HpCDF PeCDD HpCDD Congeners

Figure 42. Dioxins/furans in Bayou Grande seafood tissues.

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Figure 43 compares dioxin/furan profiles of sediment and crab hepatopancreas. Three of the dominant sediment congeners, OCDD (Octachlorodibenzo-p-dioxin), HpCDD (1,2,3,4,6,7,8- Heptachlorodibenzo-p-dioxin), and OCDF (Octachlorodibenzofuran) were not the dominant congeners in the crab. Instead HpCDF (1,2,3,4,6,7,8-heptachlorodibenzofuran) and HxCDF (1,2,3,6,7,8-hexachlorodibenzofuran) were dominant for upper bayou crab hepatopancreas. This suggests that crabs are either selective in the specific congeners that are incorporated into their tissues or that dechlorination is occurring with these congeners once incorporated into the crab (Figure 43). The crab tissues have small amounts of the most toxic congener, TCDD (2,3,7,8- tetrachlorodibenzo-p-dioxin), that was not detected in Bayou Grande sediments. There is no obvious explanation for the presence of TCDD in crabs and not in the surrounding sediments. Speculation leads to several possible hypotheses: the detection of TCDD in crabs and not in sediments might represent either dechlorination of the OCDD and other dioxins to TCDD in the crab, or crabs may be consuming other sedimentary organisms that have accumulated TCDD from other sediment levels or sources, or the bioaccumulation of undetected sediment TCDD to detectable levels in the crab.

Dioxins/furans in Crab Hepatopancreas and Sediment 100 Crab Hepatopancreas 90 139 ng HpCDD 1277 ng OCCD Sediment 80

ng mass/kg wt ng mass/kg 30

0 TCDF TCDD OCDF OCDD HxCDF HxCDF HxCDF HxCDF HxCDD HxCDD HxCDD PeCDF PeCDF HpCDF HpCDF PeCDD HpCDD Congeners

Figure 43. Dioxins/furans in Bayou Grande crab hepatopancreas and sediment.

Figure 44 represents dioxin/furan congener concentrations from oyster tissues and sediments. The sediment congener concentrations have been reduced by a factor of 100 to make the profiles more readily comparable visually. While not identical it does appear that overall profiles are much more similar than was the case for crab tissues. It appears that incorporation of dioxin/furan congeners in oysters, but not the crab hepatopancreas, is related to the relative proportions of the congeners in the sediments.

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Dioxins/furans in Bayou Grande Sediment/100 & Oyster

16 Oyster 14 Sediment ng conc/100

6 ng mass/kg wt wt ng mass/kg 4

0 TCDF TCDD OCDF OCDD HxCDF HxCDF HxCDF HxCDF PeCDF PeCDF HpCDF HpCDF HxCDD HxCDD HxCDD PeCDD HpCDD Congeners

Figure 44. Dioxins/furans in Bayou Grande oysters and sediment.

VII.2.12 PCBs in Bayou Grande Sediments and Seafood Tissues Dioxin-like PCBs in sediments can enter into the food chain and upon accumulating in seafood are consumed by humans with probable impacts upon human health. In Bayou Grande dioxin- like PCB profiles for seafood closely reflect the profiles for sediments (Table 43, Figure 45). In general, the seafood tissue profiles for dioxin-like PCBs better reflect the sediment profiles than was the case for dioxins/furans.

Table 43. Dioxin-Like PCBs in crab and oyster tissues [ng/kg wet wt] and sediment [ng/kg dry weight]. Upper bayou Lower bayou Upper Lower bayou congener crab crab bayou crab Oyster Sediment crab muscle hepatopancreas hepatopancreas muscle PCB-77 136 224 91.8 45.3 34.5 76.88 PCB-81 14 27.1 7.59 4.25 2.1 9.43 PCB-105 2,250 3,070 0.64 0.6 206 448.14 PCB-114 138 162 5.09 2.51 6.4 18.73 PCB-118 12,000 13,500 426 225 1130 1,121.46 PCB-123 226 232 7.24 3.75 15.3 5.13 PCB-126 73.4 213 3.03 3.99 30.8 31.83 PCB- 2,141 3,160 65 41 164.1 304.61 156/157 PCB-167 1,650 2,460 51.5 34 150 141.21 PCB-169 19.5 4.9 0.81 1.18 7.0 31.76 PCB-189 270 459 6.02 3.44 14.7 59.97 Sum 18,917.9 23,512 664.72 365.02 1,760 2,249.16

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Dioxin-like PCBs in Seafood Tissues & Sediment

16000 Upper Crabs Hepato

14000 Lower Crabs Hepato Upper CrabMuscle

12000 Lower Crab Muscle Oyster 10000 Sediment

8000

6000 ng mass/kg wt ng mass/kg

4000

2000

0 77 81 105 114 118 123 126 167 169 189 Congeners 156/157 Figure 45. Dioxin-like PCBs in seafood tissues and sediment.

In Figure 46 and Figure 47 are displayed the dioxin-like PCB profiles of lower bayou crab hepatopancreas, oyster tissues, and sediment. The concentration of the sediment in Figure 46 has been multiplied by a factor of ten to make the data visually more comparable. It is readily apparent that the crab hepatopancreas and oyster tissues have profiles that are very similar to that of sediments for the congeners present at the highest concentration (PCB-105, PCB-118, PCBs- 156&157, PCB-167, and PCB 189). It appears that dioxin-like PCBs are incorporated proportionally relative to their sediment concentrations into the sample seafood tissues. Similar findings were observed for dioxin-like PCBs for tissues and sediments in Bayou Chico (Mohrherr et al., 2006).

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Dioxin-Like PCBs in Crab Hepatopancreas and Sediment

16000 Crab Hepatopancreas

14000 Sediment * 10

12000

10000

8000

6000 ng mass/kg wt ng mass/kg

4000

2000

0 77 81 105 114 118 123 126 167 169 189

Congeners 156&157

Figure 46. Dioxin-like PCBs in crab hepatopancreas tissue and sediment.

Dioxin-Like PCBs in Oyster Tissues and Sediment

1200 Oyster

1000 Sediment

800

600

400 ng mass/kg wt ng mass/kg

200

0 77 81 105 114 118 123 126 167 169 189

Congeners 156&157

Figure 47. Dioxin-like PCBs in Bayou Grande oyster tissue and sediment.

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VII.3 Pesticides

Only five detections of organochlorinated pesticides occurred in the 23 sediment grab samples from Bayou Grande (Table 44). These five detections occurred in only two samples, GF-21 and GF-22. Two of the detections were for alpha-BHC, one for delta-BHC and one for 4,4'-DDT. The 4,4'-DDT detection of 3.8 ug/kg was above the TEL of 1.19 ug/kg. For alpha-BHC and delta-BHC there are no FDEP SQAGs. There were more detections in the 1990’s (Table 5, Table 6) (DeBusk et al., 2002; EnSafe, 1998; 2003; 2004)). The data for samples taken in the 1990’s covering a similar set of pesticides using similar methodology showed a detection rate of about 37% as compared to 0.09% in the current study. Most of these pesticides have not been applied for many years and appear to be declining in sediment concentrations and rate of detection. Sediment transport into and out of the Bayou can reduce the concentrations of the pesticides as can abiotic degradation and biodegradation. There were no detections in the water column for any pesticides (Table 45) which suggests that currently that there is no transport of organochlorinated pesticides into Bayou Grande from surface sources.

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Table 44. Organochlorinated pesticide concentrations in surface sediments [ug/kg]. Analyte GF-1 GF-2 GF-3 GF-4 GF-5 GF-6 GF-7 GF-8 GF-9 GF-10 GF-11 alpha-BHC <0.931 <1.4 <1.6 <2.1 <2.9 <2.6 <3 <2.7 <2.6 <0.23 <0.23 gamma-BHC <3.1 <<4.6 <5.3 <6.7 <9.4 8.4 <9.7 <8.8 <8.5 <0.73 <0.74 (Lindane) beta-BHC <1.4 <2.1 <2.4 <3.1 <4.3 <3.9 <4.5 <4.1 <3.9 <0.34 <0.34 delta-BHC <0.72 <1.1 <1.3 <1.6 <2.3 <2 <2.3 <2.1 <2.1 <0.18 <0.18 Heptachlor <2.5 <3.7 <4.2 <5.4 <7.5 <6.7 <7.7 <7 <6.8 <0.58 <0.59 Aldrin <1.1 <1.7 <1.9 <2.4 <3.4 <3 <3.5 <3.2 <3.1 <0.26 <0.27 Heptachlor <0.83 <1.3 <1.5 <1.9 <2.6 <2.3 <2.7 <2.4 <2.3 <0.2 <0.2 Epoxide gamma- <3.8 <<5.7 <6.5 <8.3 <12 <11 <12 <11 <11 <0.9 <0.91 Chlordane alpha-Chlordane <0.72 <1.1 <1.3 <1.6 <2.3 <2 <2.3 <2.1 <2.1 <0.18 <0.18 4,4’-DDE <0.72 <1.1 <1.3 <1.6 <2.3 <2 <2.3 <2.1 <2.1 <0.18 <0.18 Endosulfan I <0.88 <1.4 <1.6 <2 <2.7 <2.5 <2.8 <2.6 <2.5 <0.21 <0.22 Dieldrin <1.4 <2.1 <2.4 <3.1 <4.3 <3.9 <4.5 <4.1 <3.9 <0.34 <0.34 Endrin <1.1 <1.6 <1.8 <2.3 <3.2 <2.9 <3.3 <3 <2.9 <0.25 <0.25 4,4’-DDD <2.4 <3.6 <4.1 <5.3 <7.4 <6.6 <7.6 <6.9 <6.7 <0.57 <0.58 Endosulfan II <1.1 <1.7 <1.9 <2.4 <3.4 <3 <3.5 <3.2 <3.1 <0.26 <0.27 4,4’-DDT <0.77 <1.2 <1.4 <1.7 <2.4 <2.2 <2.5 <2.3 <2.2 <0.19 <0.19 Endrin Aldehyde <1.1 <1.6 <1.8 <2.3 <3.2 <2.9 <3.3 <3 <2.9 <0.25 <0.25 Methoxychlor <0.83 <1.3 <1.5 <1.9 <2.6 <2.3 <2.7 <2.4 <2.3 <0.2 <0.2 Endosulfan <0.98 <1.5 <1.7 <2.2 <3.1 <2.8 <3.1 <2.9 <2.8 <0.24 <0.24 Sulfate Endrin Ketone <1.6 <2.4 <2.8 <3.6 <5 <4.5 <5.1 <4.7 <4.5 <0.39 <0.39 Toxaphene <88 <140 <160 <200 <270 <250 <280 <260 <250 <21 <22

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Table 44. Organochlorinated pesticides in surface sediments [ug/kg] (continued). Analyte GF-12 GF-13 GF-14 GF-15 GF-16 GF-17 GF-18 GF-19 GF-20 GF-21 GF-22 GF-23 alpha-BHC <0.23 <3.1 <3.7 <2.6 <0.71 <3.5 <0.86 <0.28 <1.1 4.7 i 8.1 i <3.7 gamma-BHC <0.74 <11 <13 <8.4 <2.4 <12 <2.8 <0.91 <3.5 <9.4 <12 <12 (Lindane) beta-BHC <0.34 <4.6 <5.6 <3.9 <1.1 <5.2 <1.3 <0.42 <1.6 <4.3 <5.1 <5.5 delta-BHC <0.18 <2.4 <2.9 <2 <0.56 <2.7 <0.67 <0.22 <0.81 <2.3 7.5 i <2.9 Heptachlor <0.59 <8 <9.6 <6.7 <1.9 <9 <2.3 <0.73 <2.8 <7.5 <8.9 <9.5 Aldrin <0.27 <3.6 <4.3 <3 <0.83 <4 <1 <0.33 <1.3 <3.4 <4 <4.3 Heptachlor <0.2 <2.8 <3.3 <2.3 <0.63 <3.1 <0.76 <0.25 <0.93 <2.6 <3.1 <3.3 Epoxide gamma- <0.92 <13 <15 <11 <2.9 <14 <3.5 <1.2 <4.3 <12 <14 <15 Chlordane alpha-Chlordane <0.18 <2.4 <2.9 <2 <0.56 <2.7 <0.67 <0.22 <0.81 <2.3 <2.7 <2.9 4,4'-DDE <0.18 <2.4 <2.9 <2 <0.56 <2.7 <0.67 <0.22 <0.81 <2.3 <2.7 <2.9 Endosulfan I <0.22 <2.9 <3.5 <2.4 <0.67 <3.3 <0.81 <0.27 <0.99 <2.7 <3.2 <3.5 Dieldrin <0.34 <4.6 <5.6 <3.9 <1.1 <5.2 <1.3 <0.42 <1.6 <4.3 <5.1 <5.5 Endrin <0.25 <3.4 <4.1 <2.9 <0.79 <3.8 <0.95 <0.31 <1.2 <3.2 <3.8 <4.1 4,4'-DDD <0.58 <7.8 <9.4 <6.5 <1.9 <8.8 <2.2 <0.71 <2.7 <7.3 <8.7 <9.3 Endosulfan II <0.27 <3.6 <4.3 <3 <0.83 <4 <1 <0.33 <1.3 <3.4 <4 <4.3 4,4'-DDT <0.19 <2.6 <3.1 <2.2 <0.6 <2.9 <0.71 <0.24 <0.87 <2.4 3.8 i <3.1 Endrin Aldehyde <0.25 <3.4 <4.1 <2.9 <0.79 <3.8 <0.95 <0.31 <1.2 <3.2 <3.8 <4.1 Methoxychlor <0.2 <2.8 <3.3 <2.3 <0.63 <3.1 <0.76 <0.25 <0.93 <2.6 <3.1 <3.3 Endosulfan <0.24 <3.3 <3.9 <2.7 <0.75 <3.7 <0.9 <0.3 <1.1 <3.1 <3.6 <3.9 Sulfate Endrin Ketone <0.39 <5.3 <6.4 <4.4 <1.3 <5.9 <1.5 <0.48 <1.8 <4.9 <5.9 <6.3 Toxaphene <22 <290 <350 <240 <67 <330 <81 <27 <99 <270 <320 <350

1: the < indicates a nondetect showing that the result is below the method detection limit indicated by the number following.

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Table 45. Organochlorinated pesticide concentrations in water samples [ug/l]. Analyte GW-1 GW-2 GW-3 GW-4 GW-5 GW-6 GW-7 GW-8 alpha-BHC <0.0082 <0.0082 <0.0082 <0.0082 <0.0082 <0.0081 <0.0081 <0.0081 gamma-BHC <0.0085 <0.0085 <0.0085 <0.0085 <0.0085 <0.0084 <0.0084 <0.0084 (Lindane) beta-BHC <0.0088 <0.0088 <0.0088 <0.0088 <0.0088 <0.0087 <0.0087 <0.0087 delta-BHC <0.012 <0.012 <0.012 <0.012 <0.012 <0.012 <0.012 <0.012 Heptachlor <0.0099 <0.0099 <0.0099 <0.0099 <0.0099 <0.0098 <0.0098 <0.0098 Aldrin <0.0071 <0.0071 <0.0071 <0.0071 <0.0071 <0.007 <0.007 <0.007 Heptachlor <0.0082 <0.0082 <0.0082 <0.0082 <0.0082 <0.0081 <0.0081 <0.0081 Epoxide gamma- <0.0078 <0.0078 <0.0078 <0.0078 <0.0078 <0.0077 <0.0077 <0.0077 Chlordane alpha-Chlordane <0.0069 <0.0069 <0.0069 <0.0069 <0.0069 <0.0068 <0.0068 <0.0068 4,4'-DDE <0.0087 <0.0087 <0.0087 <0.0087 <0.0087 <0.0086 <0.0086 <0.0086 Endosulfan I <0.0092 <0.0092 <0.0092 <0.0092 <0.0092 <0.0091 <0.0091 <0.0091 Dieldrin <0.0076 <0.0076 <0.0076 <0.0076 <0.0076 <0.0075 <0.0075 <0.0075 Endrin <0.0093 <0.0093 <0.0093 <0.0093 <0.0093 <0.0092 <0.0092 <0.0092 4,4'-DDD <0.0082 <0.0082 <0.0082 <0.0082 <0.0082 <0.0081 <0.0081 <0.0081 Endosulfan II <0.0066 <0.0066 <0.0066 <0.0066 <0.0066 <0.0066 <0.0066 <0.0066 4,4'-DDT <0.014 <0.014 <0.014 <0.014 <0.014 <0.014 <0.014 <0.014 Endrin Aldehyde <0.0088 <0.0088 <0.0088 <0.0088 <0.0088 <0.0087 <0.0087 <0.0087 Methoxychlor <0.012 <0.012 <0.012 <0.012 <0.012 <0.012 <0.012 <0.012 Endosulfan <0.0095 <0.0095 <0.0095 <0.0095 <0.0095 <0.0094 <0.0094 <0.0094 Sulfate Endrin Ketone <0.0055 <0.0055 <0.0055 <0.0055 <0.0055 <0.0055 <0.0055 <0.0055 Toxaphene <0.52 <0.52 <0.52 <0.52 <0.52 <0.52 <0.52 <0.52 1: the < indicates a nondetect showing that the result is below the method detection limit indicated by the number following.

VII.4 Trace Metals

VII.4.1 Trace Metal Concentrations in Surface Sediments A total of 78 surface grabs were analyzed for the following trace metals: aluminum (Al), arsenic (As), cadmium (Cd), calcium (Ca), chromium (Cr), copper (Cu), iron (Fe), mercury (Hg), lead (Pb), magnesium (Mg), nickel (Ni), selenium (Se), tin (Sn) and zinc (Zn). Eight of these metals (As, Cd, Cr, Cu, Hg, Pb, Ni, and Zn) have been assigned a TEL and a PEL for FDEP SQAGs (McDonald, 1994 alb).

For arsenic there were 53 detections out of 78 samples with an average of 7.48 mg/kg for the detections (Table 46). The concentrations ranged from 0.3 mg/kg (GF-2) to 20 mg/kg (GF-21) with 22 values above the TEL (7.24 mg/kg) and no values above the PEL (41.6 mg/kg). The TEL

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was exceeded in samples located in Davenport Bayou, Navy Point Bayou, and other embayments (Figure 48). On the NAS side of the Bayou the TEL was exceeded in samples located in the Yacht Basin, Woolsey Bayou, and Redoubt Bayou. It appears that arsenic is a ubiquitous constituent of bayou sediments that often exceeds the TEL. Its presence on both sides of the bayou suggests multiple origins. It is a metal that does occur naturally, but is also present in herbicides that were formerly in use, and in atmospheric deposition. Its spatial distribution is at least in part influenced by the composition of the sediments as the correlation coefficients with clay content and organic matter content are 0.6 and 0.7 respectively.

Figure 48. Arsenic in Bayou Grande sediments. (The round symbols indicate samples of the GBc series and triangles indicate samples of the GF series.)

Cadmium was detected in 32 out of 78 samples with concentrations ranging from 0.48 mg/kg to 45 mg/kg (Table 46). The TEL for Cd (0.676 mg/kg) was exceeded by 21 samples and the PEL (4.21 mg/kg) by seven samples (Figure 49). Two of the concentrations were many times higher than the mean concentration for detections of 4.46 mg/kg (GBc-47 with 31 mg/kg in the Yacht Basin and GBc-44 with 45 mg/kg in Woolsey Bayou). Two of the other values that exceed the PEL (samples GF-2 and GF-3) are also from these two embayments. These data suggests that the two hot spots for Cd are on the NAS side of Bayou Grande but there are high concentrations of Cd elsewhere in the Bayou (Figure 49). Clay content and organic matter content do not

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significantly affect the spatial distribution of cadmium as indicated by very low correlation coefficients between these sediment characteristics and Cd concentration (0.1 and 0.3 respectively).

Figure 49. Cadmium in Bayou Grande sediments. (The round symbols indicate samples of the GBc series and triangles indicate samples of the GF series.)

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Table 46. Trace metals in surface sediments [mg/kg]. ID Al As Cd Ca Cr Cu Fe Pb Mg Hg Ni Se Sn Zn TEL NA 7.24 0.676 NA 52.3 18.7 NA 30.2 NA 0.13 15.9 NA NA 124 PEL NA 41.6 4.21 NA 160 108 NA 112 NA 0.696 42.8 NA NA 271 GBc-1 360 <0.183 <0.21 150 -8 0.35 i 220 0.73 300 <0.0052 <0.87 <0.87 <6.3 4.1 j4 GBc-2 470 0.3 <0.21 100 - 0.47 i 150 1.1 230 <0.0053 <0.88 <0.88 <6.3 2.9 i5 GBc-3 440 <0.2 <0.22 120 - 0.48 i 240 5.1 290 <0.0055 <0.93 <0.93 <6.3 3 i GBc-4 900 1.8 <0.23 1500 - 4.9 1100 4.3 470 <0.0057 <0.96 <0.96 <6.6 12 GBc-5 5001 <0.46 <0.33 90 i 1.8 0.62 i 320 1.1 250 <0.0052 0.28 i <0.95 0.78 ib 1.8 ib6 GBc-6 3500 1.7 <0.38 9000 7.1 4.2 2800 12 1000 0.0088 i 1.6 <1.1 <0.68 27 GBc-7 10000 17 1.2 i 3200 98 29 25000 64 6100 0.079 12 <2.5 2.6 ib 130 GBc-8 1500 1.3 <0.37 590 5.4 2.9 1500 3.7 530 <0.0058 0.73 i <1.1 0.87 ib 8.5 GBc-9 17000 18 0.95 i 1800 50 71 11000 44 3000 0.13 8.9 <2.5 5.2 i 130 GBc-10 12000 9.6 0.72 i 1700 50 25 12000 35 3000 0.052 i 7.8 <2 1.2 ib 87 GBc-11 14000 11 0.74 i 2000 44 22 17000 39 4100 0.06 6.9 <2.1 2.4 ib 99 GBc-12 1100 0.92 <0.37 460 3.9 1.7 1400 2.8 130 0.0066 i 0.54 i <1.1 2.6 i 7.7 GBc-13 3100 2.4 <0.45 1900 8.4 4.4 3500 7.2 290 0.016 i 1.7 <1.3 2.8 i 17 GBc-14 11000 11 0.83 i 2500 50 20 22000 34 1300 0.09 8.3 <2.3 5.9 i 91 GBc-15 12000 5.5 <0.8 2100 22 18 13000 39 840 0.07 5 <2.3 5.8 i 77 GBc-16 11000 6 <0.69 1700 20 11 12000 23 720 0.036 i 4.6 <2 5.2 i 43 GBc-17 910 <0.55 <0.39 280 1.1 i 0.75 i 810 2.2 100 <0.0056 0.34 i <1.1 2.7 i 4.7 i GBc-18 610 <0.52 <0.36 130 0.97 i 0.48 i 360 0.98 67 <0.0057 0.26 i <1.1 2.4 i 1.6 i GBc-19 1800 1.5 <0.46 670 3.2 2.9 2400 7.7 1000 0.015 i 1.1 i <1.3 1.7 ib 24 GBc-20 8100 6.3 <0.66 1800 19 11 9500 21 3100 0.013 i 3.7 <1.9 2.9 ib 38 GBc-21 1500 <0.57 <0.4 260 2.8 1.4 840 3.5 510 <0.0056 0.58 i <1.2 1.7 ib 5.8 i GBc-22 3600 3.1 <0.44 720 12 4.8 4900 8.4 1500 0.014 i 1.8 <1.3 2 ib 25 GBc-23 1000 0.99 <0.39 610 3.7 41 1400 2.8 600 0.0083 i 0.55 i <1.1 1.6 ib 8.4 GBc-24 640 <0.54 <0.38 170 1.6 0.64 i 530 1.2 370 <0.0055 0.3 i 1.1 1.6 ib 3.1 i GBc-25 540 <0.51 <0.36 150 2.2 0.72 i 550 1.4 370 0.0058 0.3 i 1 1.7 ib 3.1 i GBc-26 380 <0.5 <0.35 160 2.2 1 i 610 2.3 370 <0.0053 0.29 i <1 2 i 5.9 i GBc-27 670 0.61 i <0.37 580 2.2 1 i 750 1.8 380 <0.0054 0.24 i <1.1 2 i 4 i GBc-28 160 <0.5 <0.35 120 i 1.1 i 0.49 i 250 1 330 <0.005 <0.19 <1 2.1 i 1.4 i GBc-29 270 <0.52 <0.37 4800 2 0.88 i 670 2.2 420 <0.0049 0.24 i <1.1 <0.66 3.5 i GBc-30 120 <0.51 <0.36 100 i 0.76 i <0.37 160 0.76 300 <0.0054 <0.2 <1 2.1 i 1.1 GBc-31 310 <0.52 <0.37 140 1.5 1.2 i 280 1.4 410 <0.0052 <0.20 <1.1 <0.66 1.7 i

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Table 46. Trace metals in surface sediments [mg/kg] (continued). ID Al As Cd Ca Cr Cu Fe Pb Mg Hg Ni Se Sn Zn GBc-32 3300 3.1 0.48 i 540 18 5.4 4600 13 1400 0.024 i 2 <1.2 <0.74 26 GBc-33 810 0.58 i <0.38 510 3.1 1.1 i 1300 2.7 1000 <0.0054 0.48 i <1.1 <0.68 5.4 i GBc-34 660 <0.53 <0.38 370 3.1 0.99 i 1100 2.5 920 0.0064 i 0.51 i <1.1 <0.67 4.8 i GBc-35 8500 4.9 1.1 1200 33 11 12,000 27 2700 0.088 4 <1.8 <1.1 51 GBc-36 20000 9.7 3.1 3100 92 31 23000 91 8300 0.12 11 <3.5 2.6 i 160 GBc-37 410 0.53 i <0.36 210 1.9 0.61 i 620 1.6 490 <0.0058 0.25 i <1.0 <0.68 2.0 i GBC-38 440 <0.53 <0.37 1500 1.8 0.61 i 400 1.5 330 <0.0058 0.3 i <1.1 <0.67 1.9 i GBC-39 340 0.53 i <0.37 140 1.5 0.48 i 660 1.1 270 <0.0053 0.23 i <1.1 1.2 ib 2.1 i GBC-40 430 <0.54 <0.38 140 2 5.1 390 13 330 <0.0054 0.26 i <1.1 0.83 ib 3.5 i GBC-41 400 <0.52 <0.37 320 1.7 1.5 410 5 320 <0.0054 0.31 i <1.1 <0.66 3.3 i GBc-42 300 <0.51 <0.36 770 2.5 1 i 290 2.7 300 <0.0053 0.31 i <1 <0.65 3.7 i GBc-43 330 <0.52 <0.37 250 2.2 0.89 i 380 1.8 300 <0.0052 0.37 i <1.1 <0.66 4 i GBc-44 32000 18 452 3400 290 95 31000 370 8500 0.4 27 <4 7.5 i 400 GBc-45 880 0.62 0.61 i 2800 7.1 3.1 930 6.5 430 0.0093 i 0.75 i <1 <0.64 13 GBc-46 300 <0.51 <0.36 310 2.6 0.76 i 290 1.7 310 <0.0054 0.22 i <1 <0.64 3.2 i GBC-47 6800 8.2 31 5400 660 100 12000 390 3500 0.11 12 <2 17 i 440 GBC-48 970 <0.54 0.66 i 310 18 4.7 1100 8.8 490 0.012 i 0.55 i <1.1 1 11 GBC-49 3000 2.3 0.61 i 1200 15 11 3900 25 1200 0.038 2.1 <1.3 1.3 i 61 GBC-50 2400 <0.53 <0.38 19000 3.3 13 1900 16 3600 0.011 i 1.4 <1.1 <0.68 44 GBC- 790 <0.53 <0.38 7100 3.9 1.9 830 3.9 460 <0.0055 0.48 i <1.1 <0.67 8.5 51A7 GBC- 830 <0.55 <0.39 3200 3.3 1.4 910 3.1 470 <0.0058 0.38 i <1.1 <0.7 7.3 51B7 GBC- 12000 9.1 1.5 i 3100 56 76 20000 110 2200 0.083 8.8 <2.7 4.3 i 220 52A7 GBC- 11000 6.6 1.3 i 3200 52 66 19000 110 2100 0.099 7.4 <2.8 3.7 i 200 52B7 GBC-53 6000 7.2 1.1 1600 27 21 8400 42 2000 0.058 4.2 <1.6 1.9 i 98 GF-1 8400 4 <0.49 5700 14 5.3 9000 8 3500 0.033 i 4 <1.4 5.4 bi 24 GF-2 11000 5.3 5.1 1600 170 46 13000 72 3900 0.13 6.2 <2 16 bi 130 GF-3 13000 5.7 5.2 1600 94 23 14000 61 4700 0.16 7.2 <2.5 13 bi 150 GF-4 21000 9 2.4 2000 110 25 22000 73 6800 0.15 10 <2.9 14 bi 110 GF-5 36000 16 2.5 3500 160 34 38000 100 12000 0.19 17 <4.2 20 bi 220 GF-6 33000 14 2.4 3200 140 32 35000 92 11000 0.2 15 <3.9 18 bi 150 GF-7 42000 17 3.4 3800 200 37 42000 120 12000 0.2 20 <4.4 22 bi 190

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Table 46. Trace metals in surface sediments [mg/kg] (continued). ID Al As Cd Ca Cr Cu Fe Pb Mg Hg Ni Se Sn Zn GF-8 39000 17 2.1 i 3500 160 31 37000 100 11000 0.16 18 <3.8 19 bi 260 GF-9 39000 16 1.3 i 3500 130 29 37000 76 11000 0.15 16 <3.7 17 bi 160 GF-10 1300 0.86 <0.34 220 4.4 1.3 1500 2.7 500 0.0058 i 0.77 i <1.0 <4.5 bi 6.1 GF-11 740 0.54 i <0.35 170 2.8 0.67 i 840 1.8 380 <0.0051 0.43 i <1.0 <4.2 bi 36 GF-12 370 <0.49 <0.35 420 1.5 0.44 i 380 1.3 330 <0.0053 0.26 i <1.0 <4 bi 6.2 GF-13 35000 15 3.6 3800 180 39 36000 110 11000 0.21 17 <4.4 21 i 290 GF-14 34000 16 4.3 4300 170 56 37000 120 13000 0.29 17 <5.5 12 i 240 GF-15 22000 10 2.6 3100 93 40 25000 130 7200 0.24 10 <4.5 10 i 220 GF-16 1400 0.85 <.64 490 7.1 2.7 1700 5.9 940 0.015 i 0.74 i <1.3 2.1 ib 12 GF-17 27000 14 3.1 4300 110 66 29000 120 11000 0.34 15 <5.1 14 i 280 GF-18 3900 3.4 <.75 570 12 4.2 4900 7.9 1400 0.025 i 2.2 <1.3 2.6 ib 22 GF-19 2500 2.7 0.71 i 960 18 9.7 4000 17 1100 0.029 i 1.9 <.22 1.8 b 46 GF-20 4300 1.8 <0.78 930 5.8 3.2 4500 18 1200 0.02 i 1.5 i 2.5 1.4 i 9.8 GF-21 27000 20 2.4 i 3700 168 38 34000 100 10000 0.15 16 5.5 7.9 179 GF-22 27000 18 4.9 3800 209 53 34000 134 12000 0.22 16 7.4 8.9 215 GF-23 26000 19 6.1 3900 222 64 34000 135 12000 0.22 18 8.2 11 266 max. 42000 20 - 19000 660 100 42000 390 13000 0.40 27 - - 440 geo- 2707 4 - 968 13 5 3027 11 1190 0.05 2 - - 22 mean stand. 11850 6 - 2666 99 24 12803 70 3943 0.10 7 - - 101 dev.

1: Bold faced font indicates that the concentration is equal to or exceeds the FDEP TEL. 2: Italicized underlined font indicates that the concentration is equal to or exceeds the FDEP PEL. 3: < Indicates a nondetect showing that the result is below the method detection limit indicated by the number following the <. 4: J is a data qualifier and indicates that the value was estimated. 5: I is a data qualifier and indicates that the reported value is between the laboratory method detection limit and the laboratory practical quantitation limit. 6: B is a data qualifier and indicates that the indicated analyte was detected in the method blank. 7: Field duplicates of GBc-51 and GBc-52 were analyzed. 8: - Not analyzed or calculated.

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For chromium there were 74 detections from 74 samples with a mean of 55.4 mg/kg and a range from 0.76 mg/kg to 660 mg/kg. Ten concentrations were above the TEL of 52.3 mg/kg and another ten were above the PEL of 160 mg/kg (Table 46). The two highest concentrations were in the Yacht Basin and in Woolsey Bayou (660 mg/kg for GBc-47 and 290 mg/kg for GBc-44 respectively). In these embayments the TEL was exceeded in several samples (Figure 50). Also at NAS in Redoubt Bayou three samples (GF-5, GF-6 and GBc-36) out of 12 samples taken exceeded the TEL. Exceedances also occur in the general channel of Bayou Grande and on the northern shore. In the western part of the Bayou the SQAGs for chromium are not exceeded. These lower levels of chromium occur in what are presumably the least impacted areas of the Bayou.

Figure 50. Chromium in Bayou Grande sediments. (The round symbols indicate samples of the GBc series and triangles indicate samples of the GF series.)

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Copper was detected in 77 out of 78 samples with a mean concentration of 17.9 mg/kg for detected concentrations. The concentrations ranged from 0.35 mg/kg to 100 mg/kg (Table 46). Twenty seven of the samples exceeded the TEL (18.7 mg/kg) but none exceeded the PEL (108 mg/kg). The highest concentrations were detected in the Yacht Basin (GBc-47, 100 mg/kg) and Woolsey Bayou (GBc-44, 95 mg/kg) (Figure 51). In the western regions of the Bayou, adjacent to GBc-23, all embayments had at least one sample that exceeded the TEL. Most of the samples in the main channel exceeded the TEL with the notable exceptions of GF-16 and GF-1 that are closest to Pensacola Bay.

Figure 51. Copper in Bayou Grande sediments. (The round symbols indicate samples of the GBc series and triangles indicate samples of the GF series.)

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For lead there were 78 detections in 78 samples with a range from 0.73 mg/kg to 390 mg/kg and a mean of 41.4 mg/kg for detected concentrations (Table 46). SQAGs were exceeded by 27 samples from throughout the Bayou (Figure 52). Nineteen of the samples were above the TEL (30.2 mg/kg) and another eight above the PEL (112 mg/kg). Once again, the highest concentrations were in the Yacht Basin (GBc-47, 390 mg/kg) and Woolsey Bayou (GBc-44, 370 mg/kg). Correlation coefficients between Pb concentration and clay content (0.4) and organic matter content (0.6) indicate that sediment characteristics do not strongly affect the spatial distribution of Pb in Bayou Grande.

Figure 52. Lead in Bayou Grande sediments. (The round symbols indicate samples of the GBc series and triangles indicate samples of the GF series.)

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Mercury was detected in 49 out of 78 samples with a mean for detections of 0.098 mg/kg and a range from 0.0058 mg/kg to 0.4 mg/kg (Table 46). Mercury was found in concentrations exceeding the TEL (0.13 mg/kg) in 17 samples but no concentration exceeded the PEL (0.696 mg/kg) (Figure 53). Sample GBc-44 in Woolsey Bayou had the highest concentration. In general, however, the GF series of samples from the deeper regions of the Bayou had the highest concentrations, indicating that most of the sediment mercury is contained in sediments in the channel of the Bayou.

Figure 53. Mercury in Bayou Grande sediments. (The round symbols indicate samples of the GBc series and triangles indicate samples of the GF series.)

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Nickel was detected in 71 out of the 78 samples with a mean of 5.4 mg/kg for detections and a range of 0.22 mg/kg to 27 mg/kg (Table 46). A total of 10 samples had concentrations above the TEL (15.9 mg/kg) and no detections were above the PEL (42.8 mg/kg) (Figure 54). There was only one detection in the GBc series that exceeded the TEL, GBc-44 in Woolsey Bayou. This concentration was the highest concentration measured in the Bayou. Two detections from the GF series, both from Navy Point Bayou, were also above the TEL. Nickel has higher correlation coefficients with clay content (0.6) and organic matter content (0.8) indicating that the sediment composition affects the concentration and spatial distribution of nickel. These correlation coefficients are higher than the other trace metals.

Figure 54. Nickel in Bayou Grande sediments. (The round symbols indicate samples of the GBc series and triangles indicate samples of the GF series.)

Zinc was detected in all samples and ranged in concentration from 1.1 mg/kg to 440 mg/kg. The mean was 75.2 mg/kg (Table 46). The TEL (124 mg/kg) was exceeded in 17 samples and four other samples exceeded the PEL (270 mg/kg) (Figure 55). The two highest concentrations occurred in the Yacht Basin (GBc-47, 440 mg/kg) and Woolsey Bayou (GBc-44, 400 mg/kg). Zinc was detected in seven embayments above SQAGs including Redoubt Bayou, Woolsey Bayou, Davenport Bayou, and Navy Point Bayou. In five of these embayments concentrations above SQAGs were found in the most distal parts of the embayment, which is suggestive of stormwater transport. Zinc was also detected above SQAGs from GF-9 eastward in many of the

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GF series samples, indicating that sediments of the main body of Bayou Grande are also impacted by zinc.

Figure 55. Zinc in Bayou Grande sediments. (The round symbols indicate samples of the GBc series and triangles indicate samples of the GF series.)

Al, Ca, Fe, Mg, Se, and Sn do not have FDEP SQAGs. Al and Fe are stable metals that can be used to standardize the concentrations for the trace metals and to determine their anthropogenic component. Al was detected in all samples and the range varied from 120 mg/kg to 42,000 mg/kg with a mean of 8,811.3 mg/kg. Ca was detected in all samples and its range varied from 90 mg/kg to 19,000 mg/kg with a mean of 2030 mg/kg. Fe was detected in all samples and its range varied from 150 mg/kg to 42,000 mg/kg with a mean of 10083 mg/kg. Mg was detected in all samples and its range varied from 67 mg/kg to 13,000 mg/kg with a mean of 2978 mg/kg. Se was in detected in 6 out of the 78 samples and averaged 4.3 mg/kg for its six detections. Its range for detections varied from 1 mg/kg to 8.2 mg/kg. Sn was detected in 54 out of 78 samples and averaged 6.3 mg/kg for its detections with a range for detections from 0.78 mg/kg to 22 mg/kg.

VII.4.2 Origin of Trace Metals in Surface Sediments To examine if trace metals were of anthropogenic origin or associated with sediment composition we performed PCA on a trace metal dataset that included Al, Fe, and Mg. These three metals are typically of sediment origin and any other trace metal that has component

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loadings similar to those of these three metals most likely also derived from the sediments. This approach is similar to the one used to identify the potential origin of trace metals in soils and sediments elsewhere (e.g. Boruvka et al., 2005). Cadmium, selenium and tin were not included in the dataset because of the large number of below detection limit results for these metals. Results showed that two components had eigenvalues >1 and together explained 82% of the variance after varimax rotation. The third component had an eigenvalue of 0.7 and explained 10% of the variance.

A plot of the component loadings shows that Ni and As are associated with Al, Fe, and Mg, (Figure 56). This indicates that Ni and As are related to the composition of the sediment, and thus are not solely of anthropogenic origin. This finding is consistent with the relative high correlation coefficients between the concentrations of these metals and sediment characteristics (see above). Another PERCH task, examining trace metal pollution of surface soils, has also found a close relationship with parent material composition for Ni content, but not for arsenic content. Arsenic, however, has a relatively high background concentration in the region and its association with sediment composition is not unexpected. Mercury also plots relatively close to Al, Fe and Mg in the factor loading plot (Figure 56) but has a loading on the third component that is larger than Al, Fe and Mg, and also Ni and As. The PCA does not allow to draw a firm conclusion about the origin of mercury but mercury is most likely at least in part of anthropogenic origin. Copper, lead, chromium and zinc have component loadings that are very different from those of Al, Fe and Mg (Figure 56), which is a strong indication that these four metals are not of natural origin.

Factor loading plot

1.2

1 Cr

2 0.8 Zn Cu 0.6

Component 0.4 Hg Fe As Mg 0.2 Al

0 0 0.2 0.4 0.6 0.8 1 1.2 Component 1

Figure 56. Factor loading plot for first three components for trace metals in surface sediments. Size of circle represents magnitude of loading on third component.

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To further examine which areas of the Bayou underwent the strongest anthropogenic influence on trace metal concentrations we standardized the concentrations for copper, lead, chromium, zinc, and mercury, the metals that are the most likely to be of anthropogenic origin. The concentrations were standardized by dividing the concentrations of the metals by that of aluminum. When the analytical result for the metals was below the detection level we used half the detection level as the concentration. Aluminum is suitable for standardization because it is derived from the source rock or sediment and is relatively stable geochemically. Standardizing the data with Al further removes the influence sediment composition may have had on these metals. We then calculated the geomean for the standardized concentrations in an approach similar to the one used for the often employed Pollution Load Index (Madrid et al., 2002). The resulting index we call the average standardized pollution index (ASPI). Results show that three of the four sites most polluted with anthropogenic trace metals are in the Yacht Basin (GBc-47, GBc-48, and GF-2) (Table 47). The fourth site (GBc-40) is also located on the south side of the Bayou in a broad unnamed embayment between Woolsey Bayou and Redoubt Bayou (Figure 57). The least polluted sites are GF-1 in Pensacola bay near the mouth of Bayou Grande, GBc-5 in the midsection of the north shore, and several sites in the upper reaches of the Bayou near the westernmost extend of our sampling area (Table 47, Figure 57). These observations indicate that activities related to NAS are the main source of anthropogenic trace metals in Bayou Grande and that relatively unpolluted sediments are present in the Bayou near Pensacola Bay in the east and wetlands in the west. The low index in the headwaters of the Bayou in the west is opposite to what was found by PERCH for Bayou Texar, which had high overall trace metal concentrations in its headwaters (Mohrherr et al., 2005).

Table 47. Average standardized pollution index for anthropogenic trace metals. Site ID ASPI1 Site ASPI GBc-47 12.35 GBc-19 1.30 GBc-48 3.10 GF-21 1.25 GBc-40 3.07 GF-13 1.18 GF-2 2.71 GBc-3 1.18 GBc-52 2.61 GBc-39 1.17 GBc-44 2.60 GBc-35 1.16 GBc-23 2.58 GBc-37 1.16 GF-19 2.52 GBc-51 1.16 GBc-49 2.52 GBc-10 1.15 GBc-52A 2.48 GF-4 1.14 GBc-45 2.40 GBc-25 1.13 GBc-42 2.32 GBc-12 1.07 GBc-29 2.25 GBc-38 1.04 GBc-7 2.17 GBc-11 1.01 GBc-53 2.13 GBc-33 1.00 GBc-30 2.09 GF-5 0.99 GBc-28 2.07 GBc-15 0.95 GBc-46 1.97 GBc-27 0.95

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Table 47. Average standardized pollution index for anthropogenic trace metals - continued. Site ID ASPI1 Site ASPI GBc-41 1.96 GF-6 0.94 GBc-26 1.93 GBc-22 0.94 GBc-43 1.88 GF-7 0.92 GF-23 1.88 GF-8 0.91 GF-3 1.83 GBc-13 0.90 GBc-50 1.68 GF-18 0.89 GF-17 1.67 GBc-8 0.88 GF-22 1.64 GBc-1 0.87 GF-15 1.63 GBc-6 0.86 GF-16 1.54 GF-10 0.81 GF-11 1.53 GBc-24 0.75 GBc-31 1.53 GF-9 0.72 GF-12 1.44 GBc-20 0.70 GBc-51A 1.44 GBc-2 0.70 GBc-4 1.41 GBc-16 0.65 GBc-32 1.37 GF-20 0.64 GBc-34 1.34 GBc-17 0.63 GBc-9 1.33 GBc-21 0.60 GBc-14 1.32 GBc-18 0.55 GF-14 1.32 GF-1 0.48 GBc-36 1.31 GBc-5 0.08 1: Table is sorted on the ASPI, not the site ID.

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Figure 57. Spatial distribution of average standardized pollution index (ASPI) for trace metals in surface sediments. Map highlights the six sites with the highest and the six sites with the lowest ASPI.

VII.4.3 Trace Metal Concentrations in Vibracores There is an overall tendency for trace metals to have the highest concentrations at the surface (level A) (Table 48). Metals that are not primarily released by anthropogenic activities sometimes reach their highest concentrations at deeper levels. Arsenic, for example, is found throughout different depths of Bayou Grande. Arsenic is an element that is widely distributed in the Earth’s crust and is a natural constituent of soils and sediments in the northwest Florida region (Chen et al., 2002) and, as shown above, is of natural origin in the sediments of Bayou Grande. For several cores the highest concentrations of As were found at the deepest level (GV- 6C, GV-10C and GV-11C). Some of these concentrations were above the TEL and one was above the PEL. The locations of GV-6, -10 and -11 are not near any known sources of arsenic.

Detections for the metal Cd were infrequent compared to most of other metals and the highest concentrations were detected at the surface of the sediments in Woolsey and Navy Point Bayous. The infrequent and surficial occurrence of Cd suggests that Cd in Bayou Grande is of recent and likely of anthropogenic origin. The four surface detections above SQAGs were in embayments: Navy Point Bayou and Woolsey Bayou (both sites of multiple pollutants), near the mouth of a small embayment, and finally in a vibracore located in the deepest basin of the Bayou (GV-11).

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Table 48. Trace metals in vibracores [mg/kg]. Sample As Cd Cr Cu Pb Hg Ni Zn Al Ca Fe Mg TEL 7.24 0.676 52.3 18.7 30.2 0.13 15.9 124 NA NA NA NA PEL 41.6 4.21 160 108 112 0.696 42.8 271 NA NA NA NA GV-1A 16 6 203 56 184 0.23 17 266 31100 3670 35800 9910 GV-1B 7 <0.61 17 2.9 4.6 <0.01 4.8 18 6580 445 13200 1680 GV-1C 5.4 <0.55 3.1 1.1i 1.2 <0.01 1.2i 2.8i 1070 60i 3450 111 GV-2A 2.1 <0.54 8.4 2.5 5.1 0.01i 1.3 10 2620 383 3180 1060 GV-2B 1.1 <0.5 4.6 0.93i 1.3 <0.01 1.5 4.6i 2320 215 2520 890 GV-2C <0.5 <0.51 4 0.82i 1.3 0.01i 1i 3.5i 1760 196 1320 671 GV-4A 6.1 5.2 120 25 66 0.075 8.5 111 12000 1100 11000 3100 GV-4B 3.5 <0.57 13 2.7 2.9 0.011i 4 15 4700 330 5200 1400 GV-4C 1.5 <0.51 3 0.96i 0.82 <0.005 1.1i 3.6i 1700 130 1700 460 GV-4D 1.9 0.6 5.8 1.6 1.6 0.006i 2.2 6.5 2100 150 2300 660 GV-5A 0.98 <0.52 5 2.1 4 0.007i 0.87i 8.8 1600 300 1900 640 GV-5B 1.3 <0.48 5.1 0.88i 1.6 0.005i 1.5 4.9i 1800 210 2300 830 GV-5C 2.6 <0.5 6.3 1.3 2.2 0.007i 1.9 6.6 2400 240 3000 790 GV-6A 15 1.7i 95 41 61 0.13 15 140 26000 3300 23000 7100 GV-6B 5.4 <0.58 8.3 1.9 2.8 0.012i 2.7 7.1 3800 630 6100 1400 GV-6C 28 <0.48 8.7 2.4 2.8 0.01i 17 15 3000 240 12000 490 GV-7A 13 <1.2 55 27 55 0.16 12 98 24000 3000 25000 5100 GV-7B <0.48 <0.49 2.1 0.6i 1.3 <0.004 0.73i 1.9i 930 300 1800 160 GV-7C 0.87 <0.51 2.7 0.64i 1 <0.005 0.85i 1.8i 1300 290 2700 98 GV-8A 1.1 <0.53 8.3 15 4.5 0.011i 1.9 18 7400 270 1900 660 GV-8B 7.7 <0.85 17 2.8 6.3 0.014i 8 11 9800 1900 9300 3000 GV-9A <0.5 <0.52 4.1 1.1i 3 0.007i 0.41i 5.1i 660 140 930 340 GV-9B 0.6i <0.5 3.2 0.66i 1 <0.005 0.75i 3i 1100 130 1200 520 GV-9C 3.2 <0.5 6.6 1.2 1.8 <0.005 2 6.2 2300 150 4500 480 GV-9D 3.5 <0.5 8.5 1.6 2.1 <0.005 2.2 8.2 2400 150 4800 520 GV-10A 0.6 <0.49 5.8 1.5 4.3 0.005i 0.59i 6.9 970 280 1300 430 GV-10B 1.9 <0.48 5.9 1.2 1.9 <0.005 1.8 5.6i 1900 210 2600 730 GV-10C 24 <0.63 52 8.5 14 0.021i 13 53 17000 840 29000 3000 GV-11A 27 8.2 357 77 205 0.41 32 341 76100 4540 59000 14600 GV-11B 15 0.24i 75 9.6 14 0.042i 21 76 48200 1830 31300 9640 GV-11C 48 <0.18 92 13 19 0.055i 28 91 58300 2040 44600 10900 GV-11D 13 <0.16 50 9.7 11 0.039i 15 46 35000 2770 24500 7530 1: Bold faced font indicates that the concentration is equal to or exceeds the FDEP TEL. 2: Italicized underlined font indicates that the concentration is equal to or exceeds the FDEP PEL. 3: < Indicates a nondetect showing that the result is below the method detection limit indicated by the number following the <. 4: I is a data qualifier and indicates that the reported value is between the laboratory method detection limit and the laboratory practical quantitation limit.

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Chromium was found at all levels of the cores and there were no non-detects. The highest chromium concentrations were most often found at the surface. In cores GV-5, GV-8, GV-9 and GV-10, however, the highest concentrations were observed at deeper levels. There were seven concentrations that exceeded FDEP SQAGs (Table 48). The SQAG exceedances were at the surface, except for GV-11 in which all depth levels exceeded the SQAGs. Chromium is soluble in the aqueous phase. In soil chromium is mostly present as insoluble carbonate and oxide of chromium(III). The presence of chromium throughout the sediment of Bayou Grande may reflect the mobility of chromium in these saturated sediments. A recent study of chromium behavior in intertidal sediments and porewaters also found that chromium buried in sediments is relatively mobile and supports modest levels of soluble metal (Hursthouse et al., 2001).

Copper was found at all levels in the sediments, but only exceeded FDEP SQAGs at the surface in Navy Point Bayou (GV-1A), Woolsey Bayou (GV-4A), near Gulf Beach Highway Bridge (GV-6A), the western end of the Bayou (GV-7A) and one site in the channel (GV-11A). The generally higher concentrations of copper at the surface most likely result from the anthropogenic production and release of copper.

Lead was found throughout all levels of the vibracores but was generally most concentrated at the surface. The surface sediments exceeded SQAGs in five cores: Navy Point Bayou (GV-1A), Woolsey Bayou (GV-4A), near Gulf Beach Highway Bridge (GV-6A), western end of Bayou Grande (GV-7A) and the main Bayou channel (GV-11A). The generally higher concentrations of lead at the surface are likely due to anthropogenic causes.

Mercury in lower vibracore levels was often present at trace and non-detect levels. Only GV-10 had a sample from the lowest level that had a higher concentration that was higher than that at the surface. Mercury was found throughout all levels of the vibracores but was observed to exceed SQAGs for five cores at the surface: Navy Point Bayou (GV-1A), Gulf Beach Highway Bridge (GV-6A), western end of Bayou Grande (GV-7A), and the main Bayou channel (GV- 11A). Sample GV-4A from Woolsey Bayou that routinely had the highest sediment concentration was below SQAG.

Nickel and zinc were also present at all depth levels (Table 48). In vibracore GV-11 the TEL for nickel was exceeded at 0, 1, and 2 m depth. Zinc was detected above the PEL in the surface sample from GV-11 and the TEL was exceeded in the surface samples from GV-1 and GV-6. Zinc is one of the most common elements in the Earth's crust. It enters the air, water, and soil as a result of both natural processes and human activities but most enters the environment as the result of human activities. The presence of zinc at all sediment levels suggests that this metal was deposited by sedimentary processes. Marine sediments like those in Bayou Grande tend to be anoxic, resulting in immobility of the zinc (ATSDR, 2005e), thus making downward migration from surface levels unlikely.

The results for the vibracore samples are consistent with those for the surface samples. They show higher levels of trace metals in embayments, especially on the south side of the Bayou. At depth the concentrations are generally lower for the metals that are of anthropogenic origin, but occasionally the TEL is exceeded even at depth.

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VII.4.4 Trace Metal Concentrations in Water Eight water grab samples were analyzed for trace metals (Figure 20). Results were compared to water quality standards for class III water bodies according to State of Florida guidelines (Table 49). There were no detections for As, Cd, Se or Sn. None of the aqueous metal concentrations exceed FDEP water quality guidelines and many of the detected values show trace levels only. These aqueous concentrations show less variation than the sediment concentrations (section VII.4.1). Aluminum and the two alkaline earth metals Ca and Mg were detected in all samples, which is to be expected because they are normal constituents of seawater.

In Table 50 are the parameters that were recorded at the time of collection of the water samples and in Table 51 are the laboratory determined parameters. These values are unremarkable and are as expected for water parameters determined in a small estuary.

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Table 49. Aqueous metal concentrations in Bayou Grande [ug/l].

Metal standard1 GW-1 GW-2 GW-3 GW-4 GW-5 GW-6 GW-7 GW-8 Al None 0.083 0.17 0.12 0.029 0.035 0.36 i 0.37 i 0.32 i As <50 <0.00028 <0.00028 <0.00028 <0.00028 <0.00028 <0.0056 <0.0056 <0.0056 Cd <9.32 <0.00007 <0.00007 <0.00007 <0.00007 <0.00007 <0.0014 <0.0014 <0.0014 Ca None 240 250 260 240 240 290 230 230 Cr+6 <50 0.0013 ib3 0.0017 ib 0.0016 ib 0.0011 ib 0.0011 ib <0.0024 <0.0024 <0.0024 Cu <3.7 0.00085 i3 0.00084 i 0.00085 i 0.00073 i 0.0008 i <0.0058 <0.0058 <0.0058 Fe None 0.064 0.12 0.088 0.037 0.041 <0.17 <0.17 <0.17 Pb <8.5 0.00026 i 0.00041 i 0.00042 i 0.00019 i 0.00022 i <0.0024 <0.0024 <0.0024 Mg None 810 840 870 800 820 880 700 700 Hg <025 0.00042 i <0.00014 0.00036 i 0.00022 i 0.0002 i 0.00021 i 0.00019 i 0.00028 i Ni <8.3 0.0016 i 0.0016 i 0.0015 i 0.0013 i 0.0016 i <0.014 <0.014 <0.014 Se < 71 <0.00079 <0.00079 <0.00079 <0.00079 <0.00079 <0.016 <0.016 <0.016 Sn None <0.0002 <0.0002 <0.0002 <0.0002 <0.0002 <0.004 <0.004 <0.004 Zn <86. 0.0036 i 0.0033 i 0.0038 i 0.0026 i 0.0028 i <0.034 <0.034 <0.034

1: Water quality standards for class III water bodies from: State of Florida 62-302.530, Criteria for Surface Water Quality Classifications. 2: < Indicates that sample was not detected and that if present must be less than the indicate MDL (Minimum Detection Level). 3: b indicates detection of analyte in method blank. 4: i is a data qualifiers and indicates that the reported value is between the laboratory method detection limit and the laboratory practical quantitation limit.

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Table 50. Water parameters for Bayou Grande, February 2007. Depth Temp Salinity dO Conduct.1 Turbidity Site pH 2 [ft] [°C] [ppt] [mg/l] uS/cm [NTU]2 GW-1 5.1 ft 13.6 C 23.6 8.19 13.0 37.25 65.8 GW-1 0 11.2 18.7 8.06 15.1 30.19 60.1 GW-2 8.3 14.39 25.95 8.2 30.2 40.57 72.8 GW-2 0 11.6 20.97 8.11 34.3 33.52 63.5 GW-3 8.9 14.3 26.47 8.1 -- 41.27 3.2 GW-3 0 11.4 22.25 8.14 -- 35.36 2.3 GW-4 8.2 13.97 26.16 8.13 -- 40.76 3.8 GW-4 0 11.45 23.31 8.21 -- 36.89 -2.9 GW-5 7.056 13.46 25.1 8.21 -- 39.16 -1.6 GW-5 0 11.31 23.35 8.2 -- 36.98 -2.8 GW-6 9.152 12.84 26.09 8.19 -- 40.39 1220.8 GW-6 0 13.27 24.71 8.22 -- 38.83 1226.4 GW-7 6.636 13.86 21.35 8.24 -- 33.96 18.4 GW-7 0 13.7 20.55 8.22 -- 32.82 -1.4 GW-8 4.801 13.66 20.57 7.99 -- 32.76 -2.2 GW-8 0 13.64 21.12 8.14 -- 33.63 0.0

1: Conductivity. uS/cm is synonymous with mhos and is the reciprocal of resistance in ohms.

Table 51. Water parameters determined in laboratory. 1 2 4 Salinity TOC TSS Site pH [ppt] [mg/l] [mg/l] GW-1 8.1 22 3.6 15 GW-2 8 23 3.5 16 GW-3 8 24 2.9 23 GW-4 8.1 23 9.7 10 GW-5 8.1 22 2.8 15 GW-6 8.1 23 3 4.5 i GW-7 8.1 19 3.4 34 GW-8 7.9 19 3.6 18

1: TOC = total organic carbon 2: TSS = total suspended solids

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VIII CONCLUSIONS

Bayou Grande is the largest of the three bayous in the Pensacola metropolitan area but its environment has not been studied as extensively as that of the other two bayous. Previous studies have found elevated levels of trace metals, PAHs and PCBs on the southern shoreline adjacent to NAS but the existing environmental information was very scant for the channel and northern half of the Bayou. Until now this lack of systematic data resulted in a poor understanding of origins and impacts of SOCs in Bayou Grande. Even though PCBs were detected, their profiles and potential origin had not been thoroughly investigated. The presence of aviation and maritime activities at NAS suggest that petroleum contamination may be present in the Bayou, but this has not been fully addressed by previous studies. An outstanding question relates to the potential relationships existing between the elevated dioxin/furan and PCB levels encountered in seafood from the Bayou by another PERCH project and pollution of the Bayou’s water and sediments. The present study addressed these shortcomings in the knowledge of the environmental condition of Bayou Grande with a systematic approach on an extensive suite of pollutants.

TPH was detected in surface sediments in 17 % of the samples and detected concentrations were low, especially relative to what was observed in other nearby bayous. This indicates that although the potential for TPH contamination exists in Bayou Grande, its impacts may currently be small. PAHs were detected in almost all surface sediment samples and were most elevated in embayments along the southern shoreline of the Bayou. Several of the PAH species exceeded Florida sediment quality guidelines, indicating that these PAHs are likely to have a negative impact on biota in the embayments. The presence of naphthalene in the sediment samples in the main channel and in NAS embayments adjacent to the most active facility areas seem to point to petroleum fuels at NAS as a likely source for the PAHs. However, forensic ratios suggest that the bulk of the PAH species have a non-petroleum origin, and for surface sediments possible combustion and coal tar origins. Consequently, the naphthalene PAHs seem to have arrived in the sediments without the other PAH species that can be expected to be present in PAHs of petroleum origin. Groundwater transport from NAS or some other unknown source would reconcile the presence of naphthalenes and the forensic ratios.

Seventy-four percent of the samples had a combined dioxin/furan and PCB TEQ that exceeded the NOAA AET sediment quality guideline. Seven of these samples had a TEQ almost three times the NOAA AET. These observations show that dioxins/furans and PCBs reach concentrations that are likely to have adverse effects on biota in Bayou Grande. However, the TEQ values are not as high as in nearby Bayou Chico, which has a long history of industrial releases. The highest TEQs for Bayou Grande were found in the Redoubt, Woolsey, and Navy Point embayments, but exceedances occurred throughout the Bayou.

The dioxins/furans in Bayou Grande are of various origins. Samples from the deeper parts of the Bayou and its embayments have a dioxin/furan congener profile that is consistent with a PCP origin. PCP was not detected in the sediments but it can not be ruled out that treated wood with PCP has been employed along the Bayou Grande shoreline, or that PCP was used in antifouling paints for boats, and that the dioxin/furan impurities of the PCP accumulated in the sediments. Other samples have dioxin/furan congener profiles that indicate an origin from forest fires and

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burning of oil in industrial boilers. Samples from deeper sediment levels have a clearly different congener profile.

Homologue patterns of the PCBs in many of the surface sediment samples are similar to those of Aroclor 1260. Samples from the northern shoreline and northern embayments have homologue patterns consistent with a mixture of Aroclor 1260 and Aroclor 1254. Aroclor 1260 has been used at NAS, and 1254 is one of the most widespread Aroclors. These observations suggest that much of Bayou Grande may be influenced by PCBs originating from NAS and that along the northern shore, near residential areas such as Warrington, PCBs from other sources have been mixed in.

Dioxins/furans and PCBs in sediments can be taken up by seafood and, after consumption of seafood by humans, may negatively impact human health. In Bayou Grande, profiles of dioxin- like PCB congeners for crabs and oysters are very similar to profiles for sediments. Profiles of dioxin/furan congeners for oysters are also similar to profiles for sediments. This suggests that oysters are bioaccumulating dioxins/furans and PCBs from the sediments and in proportions similar to those in the sediments. In any case, this study shows that dioxins/furans and PCBs in Bayou Grande are elevated and are affecting seafood, and thus they have the potential to adversely affect human health.

Organochlorinated pesticides were detected in only 9 % of the sediment samples, and concentrations were low. Compared to studies carried out in the 1990s the number of detections and the concentrations of the pesticides have decreased. There were no detections of organochlorinated pesticides in water. These results suggest that the levels of the analyzed pesticides are declining in Bayou Grande and may be of little further concern.

Of the ten trace metals analyzed in this study As, Cr, Cu, Hg and Ni exceeded their respective TELs; the PEL was exceeded by Cd, Pb, and Zn. Because these metals exceed their TEL or PEL they can be assumed to have negative impacts on biota in the Bayou, but their concentrations were generally lower than in Bayous Chico and Texar, two other Pensacola bayous previously studied by PERCH. Spatial patterns of the trace metals point to sources at NAS and the urban area to the north of the Bayou.

Principal Component Analysis shows that Ni and As are not solely of anthropogenic origin, and to a large degree are derived from the sediments or source rock. Copper, Pb, Cr, and Zn appear to be of anthropogenic origin. An index that combines the standardized concentrations of these latter metals indicates that activities related to NAS are the main source of anthropogenic trace metals in Bayou Grande and that relatively unpolluted sediments are present in the Bayou near Pensacola Bay in the east and near the undeveloped wetlands in the west.

The present study was the first to systematically assess sediment concentrations of metals, mercury, PAHs, organochlorinated pesticides, dioxins/furans, PCBs, and total petroleum hydrocarbon for the entire Bayou Grande. Many of the pollutants examined exceed regulatory guidelines, including PAHs, dioxins/furans, PCBs and trace metals. Even though these pollutants may be unlikely to directly affect humans, because of limited direct contact of people with the sediments of Bayou Grande, they do have the potential to indirectly affect humans. A case in

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point are the elevated levels of dioxin/furan and dioxin-like PCB TEQ found in some seafood by another PERCH project. The exceedances of FDEP, USEPA, and NOAA sediment quality guidelines and tissue screening concentrations show that detrimental effects on the living environment may be also manifested in bayou biota.

The sediments show evidence of impacts from SOCs imported from the surrounding drainage basin. For the various classes of SOCs there are spatial differences over the Bayou relative to impact. The precise nature of the impacts is not identical . For example PAH analytical results for samples taken near NAS are characterized by higher concentrations of naphthalenes than are those adjacent to Warrington.

The major goal for future improvement of Bayou sediment quality should be to eliminate the import of contaminated materials through proper controls of runoff, and identification and remediation of contaminated groundwaters. There is of course no methodology that will totally prevent import since some SOCs are transported atmospherically. The programs towards these ends by NAS Pensacola and Escambia County, FL, must be continued and enhanced. These programs include stormwater controls, remediation of contaminated groundwater, and other environmental measures. With time, abiotic and biotic processes along with import of new sediments will diminish the impacts of the organic SOCs upon the sediments. This has apparently already happened with the organochlorinated pesticides. Many toxic metals unfortunately can remain bound to the sediments for long periods of time but becoming covered with new sediments over time will make the more tightly bound metals less available to surface biota.

Localized dredging is perhaps a viable solution for cases of exceptionally high metal concentrations that occur in some embayments. The major concern in the main bayou channel would be that the combined TEQ is frequently higher than the applicable SQAGs for environmental health and the levels of halogenated POPs are above the USEPA screening values in shellfish tissues. The pervasiveness of naphthalenes in many areas of the Bayou is also of concern. Naphthalenes are considered labile and once the sources are remediated or otherwise removed, their concentrations should quickly diminish in sediments through natural processes.

Relative to the future there is some evidence suggestive of degradation for dioxins/furans and the more highly chlorinated PCB congeners in the deeper sediments. However, other explanations are also possible. The more chlorinated POPs at the surface are not as subject to anaerobic degradation and would be expected to be available to biota for extended periods of time. Many sediment bound metals such as lead will likely remain bound in the sediment since degradation does not occur. It is possible that POPs and metals will be covered by sediments in the future that will hopefully be less contaminated. The most acute public health issues are related to bacteria originating form fecal sources that pose health risks to recreational users of the bayou’s waters. Fortunately, while some cost in involved, environmental release of fecal coliform bacteria originating from human sources can be minimized through the installation of appropriate sewerage systems.

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IX REFERENCES

Abernathy, S.G., Mackay, D., & McCarty, L.S. (1988). Volume fraction correlation for narcosis in aquatic organisms: The key role of partitioning. Environment Toxicology Chemistry, 7, 469-481.

ATSDR. (1994). Toxicological profile for Chlordane. US Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry.

ATSDR. (1995). Toxicological profile for Mirex and Chlordecone. US Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry.

ATSDR. (1996). Toxicological profile for Endrin. US Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry.

ATSDR. (1998). Toxicological profile for Chlorinated Dibenzo-p-dioxins. US Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry.

ATSDR. (1999). Toxicological profile for Mercury. US Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry.

ATSDR. (2000a). Toxicological profile for Polychlorinated Biphenyls (PCBs). US Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry.

ATSDR. (2000b). Toxicological profile for Endosulfan. US Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry.

ATSDR. (2000c). Toxicological profile for Chromium. US Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry.

ATSDR. (2001). Toxicological profile for Pentachlorophenol (PCP). US Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry. 316p.

ATSDR. (2002). Toxicological profile for DDT, DDE, and DDD. US Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry.

ATSDR. (2004). Toxicological profile for Copper. US Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry.

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ATSDR. (2005a). Draft toxicological profile for Arsenic. US Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry. (2005a). Draft toxicological profile for Arsenic.

ATSDR. (2005b). Draft toxicological profile for Lead. US Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry.

ATSDR. (2005c). Toxicological profile for Alpha-, Beta-, Gamma-, and Delta- hexachlorocyclohexane. US Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry.

ATSDR. (2005d). Toxicological profile for Nickel. US Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry.

ATSDR. (2005e). Toxicological Profile for Zinc. US Department of Health and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry.

ATSDR. (2006). Public Health Assessment, NAS Pensacola, Pensacola, FL. EPA Facility ID: FL9170024567. US Department of Health and Human Services, Federal Facilities Assessment Branch, Division of Health Assessment and Consultation Agency for Toxic Substances and Disease Registry.

Barkovskii, A. L. & Adriaens, P. (1996). Microbial dechlorination of historically present and freshly spiked chlorinated dioxins and diversity of dioxin-dechlorinating populations. American Society for Microbiology, Environmental Microbiology, 62(12), 4556-4562.

Bedard, D.L., & Quensen, J.F.III. (1995). Microbial reductive dechlorination of polychlorinated biphenyls. p. 127-216. In: Young L.Y. & Cerniglia, C.E. (eds.), Microbial transformation and Degradation of Toxic Organic Chemicals. Wiley-Liss Division, John Wiley & Sons, Inc., New York.

Ballschmiter, K. & Zell, M. (1980). Analysis of polychlorinated biphenyls (PCB) by glass capillary gas chromatography. Fresenius Zeitschrift fur Analytische Chemie, 302(1), 20-31.

Birch, G.F., Harrington, C., Symons, R.K., & Hunt, J.W. (2007). The source and distribution of polychlorinated dibenzo-p-dioxin and polychlorinated dibenzofurans in sediments of Port Jackson, Australia. Marine Pollution Bulletin, 54, 295-308.

Boruvka, L., Vacek, O. & Jehlicka J. (2005). Principal component analysis as a tool to indicate the origin of potentially toxic elements in soils. Geoderma, 128, 289-300.

Brenner, R.C., Magar, V.S., Ickes, J.A., Abbott, J.E., Stout, S.A., & Crecelius, E.A. (2002). Characterization and fate of PAH-contaminated sediments at the Wyckoff/Eagle Harbor Superfund Site. Environment Science Technology, 36(12), 2605-2613.

154

Burt R. (Ed.). (2004). Soil survey laboratory methods manual. (Soil survey investigations report No. 42). USDA NRCS. Washington, DC.

Butler, P.A. (1953). Summary of our knowledge of the oyster in the Gulf of Mexico. In P.S. Galtsoff (Ed.), Gulf of Mexico, its Origin, Waters and Marine Life. U.S. Fish and Wildlife Service. Fishery Bulletin, 64(1954), 479-490.

CDM Federal Programs Corporation. (2002). Draft remedial investigation/feasibility study for the Escambia Treating Company Site, Operable Unit 2, Pensacola, Florida. USEPA Region 4. 68-W5-0022.)

Chapman, P.M., Dexter, R.N., & Long, E.R. (1987). Synoptic measures of sediment contamination, toxicity and infaunal community composition (the Sediment Quality Triad) in San Francisco Bay. Marine Ecology, 37, 75-96.

Chatzikosma, D.G., & Voudrias, E.A. (2007). Simulation of polychlorinated biphenyls transport in the vadose zone. Environmental Geology, 53, 211-220.

Chen, C.M., Ma, L.Q., & Harris, W.G. (2002). Arsenic concentrations in Florida surface soils: Influence of soil type and properties. Soil Science Society of America Journal, 66, 632- 640.

Coates, J.D., Woodward, J., Allen, J., Philp, P., & Lovely, D.R. (1997). Anaerobic degradation of polycyclic aromatic hydrocarbons and alkanes in petroleum-contaminated marine harbor sediments. American Society for Microbiology, Applied and Environmental Microbiology, 63(9), 3589-3593.

Collard, S.B. (1991). The Pensacola Bay System biological trends and current status. Havana, FL: Northwest Florida Water Management District.

DeBusk, W.F., Poyer, I., & Herzfeld, L. (2002). Sediment quality in the Pensacola Bay System. (Technical File Report 02-03). Havana, FL: Northwest Florida Water Management District.

De Sylva, D. (1955). Report on Pollution and Fish Mortality in Bayou Chico, Pensacola, Florida. Coral Gables, FL: University of Miami Marine Lab.

Droubay, M., Lewis, J., & House, J. (1999). A Report to Escambia County Neighborhood and Environmental Services Department, Pensacola. Retrieved from the Environmental Studies Department, University of West Florida http://www.uwf.edu/GIS/Wetlands/finalreport1.htm

Dunbar J., Lin, C., Vergucht, I., Wong, J., & Durant, J. (2001). Estimating the contribution of mobile sources of PAH to urban air using real-time PAH monitoring. Science of the Total Environment, 279, 1-19.

155

Ecology and Environment, Inc. (1989). Draft contamination assessment/remedial activities investigation work plans (Groups A, B, C, D, E, F, G, J, K, M and N). NAS Pensacola, Pensacola, FL: Department of the Navy, Southern Division, Naval Facilities Engineering Command.

Elmendorf, C.E., Haith, G.S., Douglas, & Prince, R.C. (1994). Relative rates of biodegradation of substituted polycyclic aromatic hydrocarbons. In Dr. Robert E. Hinchee, Lewis Semprini, and Say Kee Ong (Eds.), Bioremediation of Chlorinated and Polycyclic Aromatic Hydrocarbon Compounds. Boca Raton: Lewis Publishers.

EnSafe/Allen & Hoshall. (September, 1995). Final remedial investigation feasibility study work plan, and final RI/FS sampling and analysis plan for sites 40 and 42 Bayou Grande and Pensacola Bay. NAS Pensacola, Pensacola, FL.

EnSafe/Allen & Hoshall. (May 22, 1997). Remedial Investigation Report: Site 42, Pensacola Bay. NAS Pensacola, Pensacola, FL.

EnSafe, Inc. (May 6, 1998). Final Record of Decision: Operable Unit 17, Site 42, Pensacola Bay. NAS Pensacola, Pensacola, FL.

EnSafe, Inc. (January 20, 1999). Final Remedial Investigation Report: Site 40. NAS Pensacola, Pensacola, FL.

EnSafe, Inc. (August 26, 2003). Final Remedial Investigation Report Vol. I, II, III: Site 40, Bayou Grande. NAS Pensacola, Pensacola, FL.

Ensafe, Inc. (January 23, 2004). Draft Record of Decision, Operable Unit 15 Site 40, Bayou Grande. Prepared for the Comprehensive Long-Term Environmental Action Navy (CLEAN). NAS Pensacola, Pensacola, FL.

Exploring Florida. (2008a). A "correct" map of Pensacola Bay showing topography of the coast, Fort Pickens, U.S. Navy Yard, and all other fortifications from the latest Government surveys in the 1860s. Map Credit: Courtesy of the Library of Congress, Geography and Map Division. http://fcit.coedu.usf.edu/Florida/maps/local/escambia/escambia.htm

Exploring Florida. (2008b). Map of Warrington and its vicinity created between 1861 and 1865 by the U.S. War Department. Map Credit: Courtesy of the National Archives and Records Administration. http://fcit.coedu.usf.edu/Florida/maps/local/escambia/m054600.htm

FDEP. (2007). Florida Healthy Beaches Program. Retrieved from http://esetappsdoh.doh.state.fl.us/irm00beachwater/default.aspx.

Federal Facilities Assessment Branch Division of Health Assessment and Consultation Agency for Toxic Substances and Disease Registry. (2003). Petitioned public health assessment, Eglin Air Force Base. (EPA Facility ID: FL8570024366). Okaloosa County, FL.

156

Federal Register. (1997). Proposed Rules. Vol. 62, No. 150. Tuesday, August 5, 1997.

Fiedler, H., Hoff, H., Tolls, J., Mertens, C., Gruber, A., & Hutzinger, O. (1994). Environmental fate of organochlorinateds in the aquatic environment. Organohalogen Compounds, 15, 199.

Fikslin, T.J. & Santoro, E.D. (2003). PCB congener distribution in estuarine water, sediment and fish samples: Implications for monitoring programs. Environmental Monitoring and Assessment, 87, 197-212.

Frame, G.M., Cochran, J.W., & Boewadt, S.S. (1996). Complete PCB congener distributions for 17 Aroclor mixtures determined by 3 HRGC systems optimized for comprehensive, quantitative, congener-specific analysis. Journal of High Resolution Chromatography, 19, 657-668.

Franks, N., & Lieb, W. (1978). Where do general anaesthetics act? Nature, 274, 339-342.

Frumkin, H. (2003). Agent Orange and cancer: An overview for clinicians. CA: A Cancer Journal for Clinicians, 53(4), 245-255.

Global Security. (2007). Retrieved from http://www.globalsecurity.org/military/facility/pensacola.htm.

Harrison, R., Smith, D.J.T., & Luhana, L. (1996). Source apportionment of atmospheric Polycyclic Aromatic Hydrocarbons collected from an urban location in Birmingham, U.K. Environment Science Technology, 30, 825-832.

Hayes, L.A., Nevin, K.P., & Lovley, D.R. (1999). Role of prior exposure on anaerobic degradation of naphthalene and phenanthrene in marine harbor sediments. Organic Geochemistry, 30, 937-945.

Hemming, J.M., Brim, M.S., & Jarvis, R.B. (2003). A survey of dioxin and furan compounds in sediments of Florida Panhandle Bay systems. Marine Pollution Bulletin, 46, 491-521.

Hemming, J.M., Brown, J.M., Brim, M.S., & Jarvis, R.B. (2005). Sediment quality survey of the Choctawhatchee Bay system in the Florida Panhandle. Marine Pollution Bulletin, 50, 889-903.

Howell, N.L., Suarez, M.P., Rifai, H.S., & Koenig, L. (2008). Concentrations of polychlorinated biphenyls (PCBs) in water, sediment, and aquatic biota in the Houston Ship Channel, Texas. Chemosphere, 70(4), 593-606.

Hursthouse, A.S., Matthews, J., Figures, J., Iqbal-Zahid, P., Davies, I.M., & Vaughan, D.H. (2001). Chromium behaviour in intertidal sediments and porewaters. UK Environmental Geochemistry and Health, 23, 253-259.

157

Karouna-Renier, N.K., Snyder, R.A., Allison, J.G., Wagner, M.G., & Rao, K.R. (2007). Accumulation of organic and inorganic contaminants in shellfish collected in estuarine waters near Pensacola, Florida: Contamination profiles and risks to human consumers. Environment and Pollution, 145, 474-488.

Lewis, M.A., Moore, J.C., Goodman, L.R., Patrick, J.M., Stanley, R.S., Roush, T.H., et al. (2001). The effects of urbanization on the chemical quality of three tidal bayous in the Gulf of Mexico. Water, air, and soil pollution, 127(1-4), 65-91.

Macauley, J.M., Summers, J.K., & Engle, V.D. (1999). Estimating the ecological condition of the estuaries of the Gulf of Mexico. Environmental Monitoring and Assessment, 57, 59- 83.

MacDonald, D.D. (1994a). Approach to the assessment of sediment quality in Florida coastal waters (Volume 1: Development and evaluation of sediment quality assessment guidelines). FDEP, Office of Water Policy. Tallahassee, FL.

MacDonald, D.D. (1994b). Approach to the assessment of sediment quality in Florida coastal waters. (Volume 2: Application of the sediment quality assessment guidelines). FDEP Office of Water Policy. Tallahassee, FL.

MacDonald, H.M. (2004). Lost–funding for stormwater management, flooding and water quality enhancement program, Escambia County, Florida.

Mackay, D., Shiu, W.Y., & Ma, K.C. (1992). Illustrated handbook of physical-chemical properties and environmental fate for organic chemicals. (Volumes I+II). Boca Raton, FL: Lewis Publishers Inc.

Madrid, L., Diaz-Barrientos, E., & Madrid, F. (2002). Distribution of heavy metal contents of urban soils in parks of Seville. Chemosphere, 49, 1301-1308.

Mohrherr, C.J., Liebens, J., Lepo, J.E., & Rao, K.R. (2005). Profiles of selected pollutants in Bayou Texar, Pensacola, Florida. Center for Environmental Diagnostics and Bioremediation, University of West Florida. Retrieved from http://www.uwf.edu/liebens/Text_final_report_Texar.pdf.

Mohrherr, C.J., Liebens, J., & Rao, K.R. (2006). Sediment and water pollution in Bayou Chico, Pensacola, Florida. Pensacola, Florida: Center for Environmental Diagnostics and Bioremediation, University of West Florida. August 11, 2006. Retrieved from http://www.uwf.edu/liebens/report_Chico_withmaps.pdf

Muller, J., Haynes, D., McLachlan, M., Bohme, F., Will, S., Shaw, G.R., Mortimer, M., Sadler, S., & Connell, D.W. (1999). PCDDS, PCDFS, PCBS and HCB in marine and estuarine sediments from Queensland, Australia. Chemosphere, 39(10), 1707-1721.

158

Murdock, J.F. (1955). An evaluation of pollution conditions in the lower Escambia River. Miami, FL: The Marine Laboratory, University of Miami Marine Fisheries Research.

NAS Pensacola. (2001). Intergrated Natural Resources Management Plan 2000-2010.

NAS Pensacola. (2003). Superfund program proposed plan. (Operable Unit 15, Site 40: Bayou Grande). NAS Pensacola Installation Restoration Program.

Naval Facilities Engineering Command, Southern Division. (2004). 2005 Site Management Plan (SMP) of the Installation Restoration Program for the NAS Pensacola. Pensacola, FL.

Nielsen, T. (1996). Traffic contribution of polycyclic aromatic hydrocarbons in the center of a large city. Atmospheric Environment, 30(20), 3481-3490.

NOAA. (1999). NOAA Screening Quick Reference Tables (SQuiRTs). Retrieved from http://response.restoration.noaa.gov/book_shelf/122_squirt_cards.pdf

Nordic Council of Ministers. (2003). Cadmium. (Cadmium Review Report no. 1, Issue no. 04).

Oleszczuk, P., & Baran, S. (2003) Degradation of individual polycyclic aromatic hydrocarbons (PAHs) in soil polluted with aircraft fuel. Polish Journal of Environmental Studies. 12(4), 431-437.

Ostrom, N., Wilson, M., & Frampton, J. (2004). Draft Lead Report. State of California Department of Toxic Substances Control, Hazardous Waste Management Program, Regulatory and Program Development Division. Sacramento, CA.

Pearce, F.P. (1989). The U.S. Navy in Pensacola: from sailing ships to naval aviation, 1825- 1930. Pensacola, FL: University Presses of Florida.

Rahman, F., Langford, K.H., Scrimshaw, M.D., & and Lester, J.N. (2001) Polybrominated diphenyl ether (PBDE) flame retardants. The Science of The Total Environment, 275(1- 3), 1-17.

Rao, K.R. (Ed.). (1978). Pentachlorophenol: chemistry, pharmacology and environmental toxicology. New York: Plenum Press.

Riisgard, H.U., & Hansen, S. (1990). Biomagnification of mercury in a marine grazing food- chain: Algal cells phaeodactylum tricornutum, mussels mytilus edulis and flounders platichthys flesus studied by means of a stepwise-reduction-CVAA method. Marine Ecology Progress Series, 62(3), 259-270.

Rostad, C., & Pereira, W.E. (1987). Creosote compounds in snails obtained from Pensacola Bay, FL, near an onshore hazardous-waste site. Chemosphere, 16(10-12), 2397-2404.

159

Schwartz, T.R., Stalling, D.L., & Rice, C.L. (1987). Are polychlorinated biphenyl residues adequately described by Aroclor mixture equivalents? Isomer-specific principal components analysis of such residues in fish and turtles. Environment Science Technology, 21(1), 72-76.

Snyder, R.A. (2006). Analysis of fecal loadings into Bayous Grande, Chico, and Texar Pensacola Bay System. Prepared for Florida Department of Health, Escambia County Health Department by Richard A. Snyder, Center for Environmental Diagnostics and Bioremediation, University of West Florida.

Sokol, R.C., Bethoney, C.M., & Rhee, G.Y. (1998). Reductive dechlorination of preexisting sediment polychlorinated biphenyls with long-term laboratory incubation. Environmental Toxicology and Chemistry, 17(6), 982-987.

Stein, E.D., Cohen, Y., & Winer, A.M. (1996). Environmental distribution and transformation of mercury compounds. Critical Reviews in Environmental Science and Technology, 1996, 26(1), 1-43.

Stout, S.A., Uhler, A.D., & Douglas, G.S. (2005). Monitoring the natural recovery of hydrocarbon-contaminated sediments with chemical fingerprinting. Environmental Claims Journal, 17 (3 & 4), 287-314.

Suarez, M.P., Rifai, H.S., Palachek, R., Dean, K., & Koenig, L. (2006). Distribution of polychlorinated dibenzo-p-dioxins and dibenzofurans in suspended sediments, dissolved phase and bottom sediment in the Houston Ship Channel. Chemosphere, 62(3), 417-429.

Tetra Tech NUS, Inc. (2003). Five Year Review, Naval Air Station Pensacola, Pensacola, Florida. (Comprehensive Long-term Environmental Action Navy contract). Naval Facilities Engineering Command. North Charleston, SC. Retrieved from http://www.epa.gov/region4/waste/npl/nplfln/pennasfl.htm

Thornburg, T. (2004). Development of SQGs for petroleum hydrocarbons: What analytes and associated SQGs should be used for bulk petroleum hydrocarbons and/or their constituents, such as PAHs? US Army Corps of Engineers - Seattle District. Retrieved from www.nws.usace.army.mil/publicmenu/DOCUMENTS/DMMO/Combined_Issue_P.

Tisdale, W.E. (1969). Report of investigations into pollution of Pensacola area waters. Northwest Florida Regional Office, Bureau of Sanitary Engineering, Florida State Board of Health.

USAF. (2006). "Turbine Fuel, Aviation, Grades JP-4 and JP-5," MIL-DTL-5624U 5 Jan 2004 and "Turbine Fuels, Aviation, Kerosene Types, NATO F-34(JP-8), NATO F-35, and JP- 8+100," MIL-DTL-83133E 1 Apr 1999. Retrieved from http://www.afcee.brooks.af.mil/pro-act/fact/petfuels.asp#4

160

US Department of Defense. (1861-1865). Map of Warrington and its vicinity created between 1861 and 1865 by the U.S. War Department.

USEPA. (1984). Mercury health effects updates: Health issue assessment, final report. (Document no. EPA 600/8-84-019F). Washington, DC: U.S. Environmental Protection Agency, Office of Health and Environmental Assessment.

USEPA. (1998). EPA’s Contaminated sediment management strategy. (EPA-823-R-98). Retrieved from http://www.epa.gov/glnpo/sediment/PCBContaminatedSedimentsStrategy.pdf.

USEPA. (2003a). Health Effects Support Document for Aldrin/Dieldrin. (EPA 822-R-03-001). USEPA, Office of Water (4304T), Health and Ecological Criteria Division. Washington, DC. Retrieved from www.epa.gov/safewater/

USEPA. (2003b). Draft dioxin reassessment draft exposure and human health reassessment of 2,3,7,8-Tetrachlorodibenzo-p-Dioxin (TCDD) and related compounds: Part III: Integrated summary and risk characterization for 2,3,7,8-Tetrachlorodibenzo-p-Dioxin (TCDD) and related compounds. (EPA/600/P-00/001Cb). Retrieved from www.epa.gov/ncea

USEPA. (2006). An Inventory of Sources and Environmental Releases of Dioxin-Like Compounds in the United States for the Years 1987, 1995, and 2000. (EPA/600/P- 03/002F). National Center for Environmental Assessment. Washington, DC.

USEPA. (2007). Tennessee NPL/NPL Caliber Cleanup Site Summaries. Retrieved from http://www.epa.gov/region4/waste/npl/nplfln/pennasfl.htm

USEPA and US Army Corps of Engineers. (2005). Site Management and Monitoring Plan (SMMP) for the Pensacola Offshore Ocean Dredged Material Disposal Site (ODMDS), Gulf of Mexico.

USGS. (2000). Pesticides in stream sediment and aquatic biota. (USGS Fact Sheet 092-00). Retrieved from http://ca.water.usgs.gov/pnsp/rep/fs09200/

USGS. (2005). Organochlorinateds in streambed sediment and aquatic biota. Retrieved from http://pubs.usgs.gov/circ/circ1171/html/organo.htm

Van den Berg, M., Birnbaum, L.S., Denison, M., De Vito, M., Farland, W., Feeley, M., et al. (2006). The 2005 World Health Organization Reevaluation of Human and Mammalian Toxic Equivalency Factors for Dioxins and Dioxin-Like Compounds. Toxicological Sciences, 93(2), 223–241.

Van Mouwerik, M., Stevens, L., Seese, M.D. & Basham, W. (1998). PAHs. In R.J. Irwin (Ed.) Environmental Contaminants Encyclopedia. National Park Service, Colorado State University.

161

Wade, T.L., Jackson, T.J., Gardinali, P.R., Chambers, L. (1997). PCDD/FCDF sediment concentration distribution: Casco Bay, Maine. Chemosphere, 34(5-7), 1359-1367.

Yunker, M.B., Macdonald, R.W., Vingarzan, R., Mitchell, R.H., Goyette, D., & Sylvestre, S. (2002). PAHs in the Fraser River Basin: a critical appraisal of PAH ratios as indicators of PAH source and composition. Organic Geochemistry, 33, 489-515.

162