Metals are naturally occurring components of the mineral part of a sediment particle. Major metals in sediments are aluminum, iron, and manganese and these are often used to differentiate types of sediment (more like terrestrial soil or limestone bedrock). Sediment composition varies naturally with local geography and environment, and so the concentrations of metals in sediments and water bodies also vary naturally. Sediments in the mainstem LSJR have widely different geologic sources. By contrast, the Cedar-Ortega system sediment characteristics suggest common geologic sources (Durell et al. 2004; Scarlatos 1993). As a result of this natural variability, it can be difficult to determine if metal levels are elevated because of human activities or simply because of the nature of the sediments. Concentrations of metals of high concern, like lead or chromium, are often compared to aluminum concentrations to try to determine what amount is the result of human input (Alexander et al. 1993; Schropp and Windom 1988). However, anthropogenic contributions of excess metals in aquatic environments are generally much greater than natural contributions (Eisler 1993).
Metals may enter aquatic systems via industrial effluent, agricultural and stormwater runoff, sewage treatment discharge, fossil fuel combustion, ore smelting and refining, mining processes, and due to leachate from metal-based antifouling paints (Reichert and Jones 1994; Kennish 1997; Evans et al. 2000; Voulvoulis et al. 2000; Echols et al. 2009). Coal and oil combustion represent a substantial release of atmospheric metals, often fated for future deposition into water bodies. Metals are only present in these fuels in small quantities; however, massive amounts of fuel are combusted. Metallic contamination also occurs with various metal-working enterprises where metal fabrications are produced and processed. Another avenue for metals to enter into aquatic environments is from leaching from hazardous waste sites (Baird 1995). Naturally occurring trace metals such as copper, zinc, and nickel are essential micronutrients required by all organisms; however, in excess, these metals, as well as non-essential metals, such as arsenic, cadmium, lead, silver, and mercury may cause adverse biological effects in aquatic organisms (Bryan and Hummerstone 1971; Dallinger and Rainbow 1993; Bury et al. 2003; Bielmyer et al. 2005a; Bielmyer et al. 2006a).
Copper and zinc are two of the most widely used elements in the world and as such are common pollutants found in freshwater and marine ecosystems (Bielmyer-Fraser et al. 2017). Copper enters aquatic systems through runoff from rivers adjacent to heavy metal mining areas (Bryan 1976); through sewage treatment discharge, industrial effluent, anti-fouling paints, refineries, as well as overflow from stormwater ponds (Guzman and Jimenez 1992; Jones 1997; Mitchelmore et al. 2003). Copper is also a constituent of several pesticides commonly used to control algae. Zinc is a major component of brass, bronze, rubber, and paint and is introduced into water systems via commercialized businesses (smelting, electroplating, fertilizers, wood preservatives, mining, etc.) and rainwater run-off (Eisler 1993). Although there are freshwater environments with only a few micrograms of zinc per liter, some industrialized areas may have problematic concentrations of over 1000 µg/L Zn (Alsop and Wood 2000). Along with copper and zinc, nickel-containing materials make major contributions to many aspects of modern life. The uses of nickel include applications in buildings and infrastructure such as stainless steel production and electroplating; chemical production, such as production of fertilizers, pesticides and fungicides; energy supply, water treatment, and coin production (Nriagu 1980; Eisler 1988b; Hoang et al. 2004). The largest use of nickel alloys and a major use of copper and zinc are in corrosion prevention. Although these applications have provided many benefits, they have resulted in increased environmental concentrations, which may have significant impact on aquatic life (Pane et al. 2003; Hoang et al. 2004). In the past, lead has also been used to a large extent in corrosion prevention, but legislation in the 1980s has limited the content of lead in paints, reduced the lead in gasoline, and eliminated the use of lead shot nationwide (Eisler 1988a). Current concerns about lead contamination in aquatic environments are mainly due to point-source discharges from mining, smelting, and refining processes, mostly for use in the production of batteries (Eisler 1988a; WHO 1995). Natural sources of lead such as erosion and atmospheric deposition from volcanoes and forest fires also contribute to the lead found in aquatic environments (WHO 1995). Elevated silver concentrations in aquatic animals occur near sewage outfalls, electroplating plants, mine waste sites, or areas where clouds have been seeded with silver iodide. The photographic industry has been the major source of anthropogenic silver discharges in the United States (Eisler 1996); however, over the last decade the use of silver, as silver nanoparticles, has substantially increased, particularly for applications in catalysis, optics, electronics, biotechnology and bioengineering, water treatment, and silver-based consumer products. Arsenic and many of its compounds are especially potent poisons, especially to insects, thereby making arsenic well suited for the preservation of wood, which has been its primary historical use.
Chromated copper arsenate, also known as CCA or Tanalith, has been used worldwide in the treatment of wood; however, its use has been discontinued in several areas because studies have shown that arsenic can leach out of the wood into the soil, potentially causing harmful effects in animals and severe poisoning in humans (Rahman et al. 2004).
Metals may be suspended in the water column for various time periods, depending on a variety of abiotic and biotic factors. In the water column, metals can reversibly bind to organic and particulate matter, form inorganic complexes, and be passed through the food chain (Di Toro et al. 2001). Various chemical reactions favor the transfer of metals through the different phases. Ultimately, metals partition in the sediment over time, as has occurred in the LSJR; however, metals may be remobilized into the interstitial water by both physical and chemical disturbances.
Metal concentrations in saltwater generally range from 0.003-16 mg/L Zn (Bruland 1980; Bruland 1983), 0.13-9.5 mg/L Cu (Kozelka and Bruland 1998), 0.2 to 130 mg/L Ni (DETR 1998; WHO 1991), and from 0.001 to 0.1 mg/L Ag (Campbell et al. 2000). The highest metal concentrations reported were measured in estuaries with significant anthropogenic inputs. However, in most cases the concentration of organic ligands, such as humic and fulvic substances, as well as the concentration of inorganic ligands exceed metal concentrations thereby forming complexes and rendering metals less bioavailable to aquatic organisms (Campbell 1995; Kramer et al. 2000; Stumm and Morgan 1996; Turner et al. 1981; Wang and Guo 2000). Aquatic animals, particularly zooplankton, have been shown to be highly sensitive to these metals (Bielmyer et al. 2006a; Jarvis et al. 2013). Lead concentrations in natural waters generally range from 0.02 to 36 µg/L, with the highest concentrations found in the sediment interstitial waters, due to the high affinity of this metal for sediment (Eisler 1988a).
Benthic biota may be affected by metals in the sediment, both by ingestion of metal-contaminated substrate and by exposure through the interstitial water. The presence of metals in the interstitial water is primarily controlled by the presence of iron sulfide in the sediments (Boothman et al. 2001). All major pollutants will displace iron and tightly bind to sulfide, thus making them less available to cause toxicity to organisms.
Once in aquatic systems, most waterborne metals exert toxicity by binding to and inhibiting enzymes on the gill or gill-like structure of aquatic organisms (Bury et al. 2003; Bielmyer et al. 2006b). This leads to a disruption in ion and water balance in the organism and ultimately death, depending on the metal concentration and exposure time. In saltwater, fish drink water to maintain water balance and therefore, the intestine is another site for metal accumulation and ion disruption (Bielmyer et al. 2005b; Shyn et al. 2012). Ingestion of metal contaminated diets can also cause intestinal metal accumulation and potentially toxicity to the consumer (Bielmyer et al. 2005b; Bielmyer and Grosell 2011; Bielmyer et al. 2012b). Decreased respiration, decreased reproductive capacity, kidney failure, neurological effects, bone fragility, mutagenesis (genetic mutation), and other effects have been observed in aquatic biota after metal exposure. Several water quality parameters can modify the toxicity of metals including: salinity, DO, dissolved organic carbon concentration (humic and fulvic substances), sulfide concentration, pH, water hardness and alkalinity, as well as other variables (Campbell 1995). The toxicity of metals may therefore vary in different parts of the LSJR, reflecting the changes in water chemistry (Ouyang et al. 2006) as well as the organisms that reside there. Metal toxicological studies using organisms or water from the LSJR are scarce. Grosell et al. 2007 and Bielmyer et al. 2013 collected Fundulus heteroclitus (killifish) from the LSJR and used them in acute (96 h) toxicological studies in the laboratory to determine the influence of salinity on copper, zinc, nickel, and cadmium toxicity to the larvae. As salinity increased, toxicity generally decreased for the metals tested. In freshwater, significant mortality to larval killifish occurred after exposure to copper (Grosell et al. 2007), zinc (Bielmyer et al. 2012a), nickel (Bielmyer et al. 2013) and cadmium (Bielmyer, unpublished work) at concentrations reported in the LSJR over the past five years (see section 2.7); however significant larval mortality was only observed after exposure to higher nickel concentrations than those found in the LSJR (Bielmyer et al. 2013). The presence of killifish is important in the LSJR because they are a common food source for many larger fish. Exposure to these metals for long time periods may cause deleterious effects, such as decreased growth and/or reproduction, in various species at even lower concentrations. Exposure to 50 µg/L for 21 days caused decreased growth in hybrid striped bass in freshwater; whereas, those exposed to the same concentration in saltwater did not suffer growth reduction (Bielmyer et al. 2006b). Generally, larval fish are more sensitive to metals than adults, and invertebrates can be even more sensitive than larval fish (Bielmyer et al. 2007). In water collected from Green Cove Springs, exposure to silver concentrations as low as 0.34 µg/L for the invertebrate crustacean, Ceriodaphnia dubia (common food sources for larval fish), and 6 µg/L for fathead minnows, respectively, caused 50% mortality to the organisms (Bielmyer et al. 2007). These silver concentrations have been reported to occur in parts of the LSJR. Many zooplankton exposed to metals, particularly through their diets, have been shown to be very sensitive to metals (Bielmyer et al. 2006a) and to accumulate metals (Bielmyer et al. 2012b). Metal exposure to the lower trophic levels may impact higher-level consumers by decreasing food availability and/or by introducing metal exposure via the diet. Sepúlveda et al. 2002 reported the accumulation of both metal and organic contaminants in the livers of Florida largemouth bass collected from four different locations in the LSJR: Welaka, Palatka, Green Cove, and Julington Creek. The highest mean liver metal concentrations were found in bass from Julington Creek (silver, arsenic, chromium, copper, zinc) and Welaka (cadmium, mercury, lead, selenium, tin). The zinc concentrations accumulated in the liver of the fish from Julington Creek were similar to those observed in adult killifish after exposure to 75 µg/L Zn in the laboratory (Shyn et al. 2012). Lead (Pb) can exist as an organometal and has a higher partition coefficient than the other metals discussed here; therefore, Pb would be preferentially distributed in more hydrophobic compartments (Eisler 1988a). Lead has been shown to exert toxic effects on a variety of aquatic organisms with sensitivity of some invertebrates as low as 4 µg/L (Grosell et al. 2006). Chronic lead toxicity in fish includes neurological and hematological dysfunctions (Davies et al. 1976; Hodson et al. 1978; Mager and Grosell 2011).
5.5.2. Current Status and Trends of Metals in Water and Sediments
184.108.40.206. Metals in Water
Generally, since 2010, a pattern of stabilized or reduced metal concentrations, particularly the maximum values, has been observed, as compared to previous years, in the LSJR mainstem. This reduction in metal concentration may reflect the recent efforts associated with TMDLs. However, the data set for metals in the water column has been substantially reduced over the years, which may contribute error in the data trend analyses. Each metal is discussed in turn below.
Arsenic With all but one exception (elevated maximum value) in 2000, the arsenic minimum, median, and maximum values in the LSJR mainstem and tributaries have been below the WQC of 50 µg/L since 1997 (Figure 5.13). Past exceedances of the WQC have mainly occurred in Cedar River, Doctors Lake, Durban Creek, and Moncrief Creek (Figure 5.20A). Mean arsenic values have decreased over time in the mainstem with the exception of a spike in 2016 (Figure 5.14). Median and maximum cadmium values in the LSJR have fluctuated since 1997 (Figure 5.15).
Cadmium Mean cadmium concentrations have generally decreased, and have been below WQC (with the assumed hardness value of 100 mg/L) in the entire LSJR since 2009 and in the mainstem since 2001 (Figure 5.16). Maximum values are now at or below WQC as well (Figure 5.15). In the past, cadmium exceedances in the tributaries have occurred specifically in Hogan Creek, as well as Cedar River, McCoy Creek, and Moncrief Creek to some degree (Figure 5.27).
Copper Copper was one of the more commonly found metals in the LSJR, based on this data set. Since 1997, maximum copper concentrations in the predominantly saltwater regions of the LSJR mainstem and tributaries have exceeded the WQC; however, since 2016 all maximum copper concentrations in the LSJR were below the WQC and within acceptable limits (Figure 5.17; 5.27). Overall, maximum copper concentrations have decreased since 2010 in the LSJR and median values have been stable (Figure 5.17). Mean copper values have significantly decreased in the saltwater regions of the LSJR mainstem since 1997 and in the freshwater regions of the mainstem since 2012 (Figure 5.18). Copper has been most problematic in the tributaries, where many exceedances have been documented (Figure 5.17; 5.27).
Lead Since 2008, maximum lead concentrations have decreased in the LSJR (Figure 5.19), with significantly decreased mean values over time in the entire LSJR, particularly the saltwater areas (Figure 5.20). In several tributaries, including Big Fishweir Creek, McCoy Creek, and Moncrief Creek, lead concentrations (median and maximum values) exceeding both freshwater and saltwater criteria have been documented, as have the maximum lead concentrations in several other tributaries; however, since 2016, all lead concentrations were below the WQC values (Figure 5.27).
Nickel Maximum nickel concentrations in the entire LSJR have decreased and remained stable since 2009, with concentrations below the saltwater and freshwater criteria of 8.3 µg/L and 52 µg/L, respectively (Figure 5.21). Additionally, mean nickel concentrations have significantly decreased over time (Figure 5.22). Since 1997, maximum nickel concentrations have been reported above WQC in several tributaries, particularly Doctors Lake, Dunns Creek, and Sixmile Creek; however, as of 2016, all nickel concentrations have been reported below WQC (Fig. 5.27).
Silver Median and maximum silver concentrations in the LSJR mainstem have fluctuated since 1997, with decreased median silver concentrations observed from 2015 until the present time (Figure 5.23). From 2006-2014, mean silver concentrations in the freshwater portion of the LSJR mainstem were elevated above the WQC of 0.07 µg/L; however, the mean silver concentrations were below the WQC since 2015 (Figure 5.24). Maximum silver concentrations within several tributaries were above both freshwater and saltwater WQC since 1997; however, as of 2016, the maximum values are at or below the WQC (Figure 5.23; 5.27).
Zinc Median and mean zinc concentrations in the entire LSJR were below the WQC and within acceptable limits since 1997, and maximum zinc concentrations were below WQC since 2008 (Figure 5.25; Figure 5.26). In the past, elevated maximum zinc concentrations were reported in Doctor’s Lake, Dunns Creek, McCoy Creek, and Butcher Pen Creek; however, current reported zinc concentrations are within acceptable limits (Figure 5.27). The metals analyzed in this report are widely used and therefore continue to enter the LSJR through point and nonpoint sources. The majority of the metal concentrations in the water column of the LSJR mainstem were at or below WQC for the last three years. The metal concentrations in the tributaries were generally the highest and therefore most problematic.
For these reasons, the current overall STATUS of metals in the water column that were evaluated in this study (including arsenic, copper, cadmium, lead, nickel, silver, and zinc) in the mainstem of the LSRJ is satisfactory with a TREND of improving. The STATUS and TREND of metal concentrations in the tributaries of the LSJR cannot be determined because of the lack of data, and is therefore uncertain.
Data Limitations It should be noted that the data set has decreased tremendously. For example, in 2007, there were 397 data points for nickel concentrations in the tributaries, and, there were only 54 data points for 2017. Additionally, these ratings are for the water column only; sediments act as a reservoir and may still contain high metal concentrations (see below). If sediments are disturbed by dredging or other activities, metals may be remobilized into the water column and may negatively impact aquatic life in the LSJR. Environment Florida’s recently released Troubled Waters report shows that US Naval Station Mayport has had more than 12 exceedances of various parameters, including nickel and copper, during a 21-month span between January 2016 and September 2017. The magnitude of potential impact is dependent on many concurring abiotic and biotic factors.
220.127.116.11. Metals in Sediments
The metals in sediments that we have evaluated in this study include arsenic, mercury, lead, cadmium, copper, silver, zinc, nickel, and chromium. Metals in general have been elevated over natural background levels in sediments all throughout the LSJR for more than two decades (Table 5.2) and continue to do so today. Many of the sediments that were analyzed since 2000 have had concentrations of chromium, zinc, lead, cadmium, or mercury (discussed in more detail below) that are greater than natural background levels (NOAA 2008), sometimes by very large amounts. Sediments in Rice Creek that were analyzed in 2002 had mercury levels that were about 100 times greater than natural background levels. High metal concentrations were found in sediments elsewhere throughout the river, including the Cedar-Ortega system, Moncrief Creek off the Trout River, Broward Creek, and Doctors Lake.
Table 5.2 Average Metal Concentrations and Percentage of Samples Exceeding Background and Sediment Quality Guidelines in the LSJR Sediments from 2000-20071 (see text in Section 5.2 for data sources)
|Average, ppm||Background, ppm1||% > Background||TEL2, ppm||% > TEL||PEL2, ppm||% > PEL|
From the 1980s to 2003, different metals exhibit slightly different trends with time, but none appear to be significantly declining in any area. Two important contributors to overall metal toxicity, zinc in the Cedar River in Area 1, and silver in Area 2, had average concentrations between their respective TELs and PELs, suggesting that the metals found throughout the LSJR individually exert a low-level stress during that time period. Metals in Area 3, the north mainstem, have increased since 1983, but the rate of increase has slowed since the mid-1990s (Figure 5.28). Although a decrease in lead concentrations were not observed from the ban of lead products from gasoline, sediment cores analyzed by other researchers give a more accurate picture of the historical record of contamination. The core studies do show recovery from lead contamination since the 1970s (Durell et al. 2005).
There is little evidence of a widespread decrease in metals between the 1980s and 2007, in contrast to the PAHs. Different metals exhibit slightly different trends with time, but none appear to be significantly declining in any area. Metals in Area 3, the north main stem, have increased since 1983, but the rate of increase has slowed since the mid-1990s (Figure 5.30). Since that time, the overall toxicity pressure from these six metals has generally remained between one and three (Figure 5.31). Although we did not see a decrease in lead concentrations from the ban of lead products from gasoline, sediment cores analyzed by other researchers give a more accurate picture of the historical record of contamination. The core studies do show recovery from lead contamination since the 1970s (Durell, et al. 2005).
For these reasons, the STATUS of metals in sediments is unsatisfactory, and the TREND is unchanged.
From 2005 to 2016, despite some hot spots, mean metal concentrations in sediments are generally present at concentrations near their TELs; however, for most metals, values above TELs have been reported (Figure 5.29). In particular, lead and mercury continue to be problematic in the LSJR (Figure 5.29). Individually, metals may exert pressure to aquatic life; however, exposure to all metals together may cause synergistic toxic effects, constituting an important class of stressor to the river. It should be noted that the number of sediment samples analyzed for metals has decreased over the past five years by more than 10-fold in some cases. In 2016, there were less than 5 samples for some metals.
For these reasons, the STATUS of metals in sediments is unsatisfactory, and the TREND is unchanged.
5.5.3. Point Sources of Metals in the LSJR Region
Most metals emitted to the atmosphere declined significantly between 2001 and 2013, with a 97% reduction in vanadium released by electric utilities accounting for much of the decline (Figures 5.30 and 5.31). In addition, zinc, nickel, copper and cobalt emissions declined significantly from 2002 to 2013 (Figure 5.30). In 2013, releases of 14 different metals to the atmosphere in the LSJR basin were reported. Zinc was the most abundant and comprised about 35% of all metal releases.
In contrast to atmospheric emissions, surface water discharges of metals increased by over 230% to a total of 71,000 pounds between 2001 and 2013. The paper industry released most total metals into the LSJR in 2013 because of the extremely large quantity of manganese that was reported (51,000 pounds). Additional metals discharged by that industry were lead (415 pounds) and mercury (0.26 pounds). Excluding manganese, electric utilities discharged about 50 times more metals than the paper industry and had more diverse effluents with 13 different metals. The metals released by electric utilities totaled 19,712 pounds in 2013 with the top five being barium, cobalt, molybdenum, nickel, and zinc.
Much of the overall increase in metals released to the LSJR is due to the electric utilities, which has had an increase of 250% in its metal discharges since 2001, despite that industry’s significant reduction in its air emissions (Figures 5.32 and 5.33). Seven of the 13 metals that were reported in 2013 by the utilities have higher release rates than in 2001. Zinc and nickel increased sharply between 2011 and 2012, while cobalt and barium increased significantly between 2007 and 2008 and have steadily increased since. Reported discharges of mercury and vanadium have decreased since 2001.
5.5.4. Mercury in the LSJR
18.104.22.168. Background: Mercury
Like most metals, mercury has natural and anthropogenic sources. As a constituent of the earth’s crust, it is released to the atmosphere by natural geologic processes. However, anthropogenic activities can substantially increase the mobilization of mercury into the atmosphere. In an assessment of national sources of mercury, the EPA determined that approximately 60% of the mercury deposited in the U.S. had anthropogenic sources (EPA 1997b). Though there is evidence there is more mercury in the atmosphere since the Industrial Revolution, there is little certainty about trends since that time (EPA 1997a).
People introduce mercury into the atmosphere by fuel combustion, ore mining, cement manufacture, solid waste incineration, or other industrial activities. Fertilizers, fungicides, and municipal solid waste also contribute to mercury loading but combustion is the primary anthropogenic source (Figure 5.34).
The LSJR emissions reflect national trends in that most waste mercury is emitted from coal power plants (EPA 1997a).
When mercury is released to the atmosphere, the most common type of release (EPA 1997a), its fate is highly dependent on the form of the mercury, meteorological conditions, and the location of the source. Elemental gaseous mercury Hg0, is the most abundant in the atmosphere and stays there for long periods of time. Oxidized species, Hg II forms, are more water-soluble and are washed out of the atmosphere and are readily transported to rivers and streams.
Local and regional modeling of the fate of mercury indicates that a substantial portion of emitted mercury travels farther than 50 km from the original source (EPA 1997a). Consequently, it is extremely difficult to isolate specific sources of mercury to a particular watershed. Considerable effort at the federal and state level has been devoted to understanding how mercury travels and cycles throughout the globe.
Once deposited into an aquatic environment, mercury can be transformed by microorganisms to an organic form, methyl mercury. Methyl mercury production is promoted by low nutrients, low oxygen, and high dissolved organic carbon levels which are typical of many Floridian lakes, blackwater streams, and wetlands. Methyl mercury binds to proteins in tissue and therefore readily bioaccumulates. All of the mercury present in prey fish is transferred to predators and the mercury biomagnifies in organisms as it travels up the food chain. High level predators with long life-spans, such as largemouth bass in freshwater and king mackerel in marine systems, accumulate the most mercury in their tissue and therefore they generally have the highest concentrations (Adams and McMichael Jr 2001; Adams et al. 2003). Humans, as top predators, consume mercury in fish also and this is the route by which most people are exposed to mercury (EPA 2001). It is important to realize that when anthropogenic mercury is mobilized to the atmosphere, it will continue to cycle, in some form, through the atmosphere, water bodies, land, or organisms (Figure 5.35).
The human health effect of mercury depends on the form, the mode of exposure, and the concentration. Methyl mercury is particularly worrisome because it is the form that is most toxic, it is most easily absorbed through the human gastrointestinal tract and it is released to the bloodstream after consumption. It passes readily into most tissues, including the brain and kidneys, where it can cause permanent damage. Exposure to pregnant women is particularly hazardous since it is passed from mothers to their children through the placenta before birth, and through nursing after birth. Methyl mercury is a neurotoxin and its effect on developing fetus’ and children is of high concern. It also appears to affect cardiovascular and immunological health of all human populations. High levels of the metallic form of mercury (Hg0) also cause problems but inorganic salts of mercury (Hg II) do not pass as easily into the brain so neural damage is not as certain (ATSDR 2000, EPA 2001).
Both the EPA and FDEP have begun to evaluate the significance of mercury contamination in water bodies based on human health risks from fish consumption, rather than based on simple water column concentrations (EPA 2001, DEP 2009a, FDOH 2016). As discussed in Section 3 of this report and below, when mercury is found in fish or shellfish, health agencies may limit consumption, particularly for women of childbearing age and children. There are 16 fresh water bodies in the LSJR basin for which the FDOH has placed consumption limits for some fish species because of mercury (FDOH 2016). In addition, there were 34 water bodies or segments of water bodies listed as impaired in the 2009 303(d) list for TMDL development based on health effects from consumption of fish contaminated with mercury (DEP 2009a) (see Section 1).
A methyl mercury fish tissue criterion has been developed that is designed to protect the health of general and sensitive populations while allowing people to consume as much fish as possible (EPA 2001, ATSDR 1999). Sensitive populations consist of children and women of childbearing age. To determine if mercury found in fish is harmful to human health, toxicologists use a reference dose (a dose that causes no ill effect) of 0.0001 mg mercury/kg human body weight per day for sensitive populations, and 0.0003 mg mercury/kg human body weight per day for the general population. These are the amounts of mercury that can be safely consumed. When fish tissue exceeds safe levels, FDOH, in concert with FWC and FDEP, issues advisories that recommend limiting consumption to a certain number of meals per week or month, or restricting it entirely. Meals should be limited for the general population when mercury in fish tissue exceeds 0.3 ppm and when it exceeds 0.1 ppm for sensitive populations. When fish tissue exceeds 1.5 ppm, the general population should not eat any of the fish. Sensitive populations should not eat any fish with mercury concentrations greater than 0.85 ppm. (EPA 2001, Goff 2010). As long as monitored fish contain low enough concentrations of mercury so that people will not consume more than the reference dose at standard rates of consumption, then no restrictions will apply.
The FL DEP issued its final report for the statewide mercury TMDL in October 2013 (see Section 1 in this report for additional information on TMDLs). The ultimate goal of the TMDL effort is to reduce the levels of mercury in fish in State waterways to safe levels where fish consumption advisories have been issued. The elements of the multi-year study to establish mercury load limits included measuring the amount of mercury that is present in Florida waterways (in fish, water and sediment), and identifying sources and fates of mercury in the State through atmospheric monitoring and modeling.
Intensive monitoring of atmospheric mercury, along with other metals and air quality parameters, was undertaken at seven sites from 2008-2010. Wet deposition of mercury was monitored at all sites and in Jacksonville, Pensacola, Tampa and Davie dry deposition was also monitored. In addition to atmospheric monitoring, extensive analysis of mercury in fish, primarily largemouth bass, and water quality was undertaken in over 100 freshwater lakes and 100 streams. The selected sites varied in acidity, trophic status and color, all parameters that were thought to affect the fate of mercury in water bodies and its uptake by fish and other organisms. These data are being used to predict levels in unmonitored sites. Mathematical models of the emissions, transport, and rates of deposition of mercury into waterways were developed as well as models to predict the concentrations in fish with different mercury loading rates and in different aquatic environments. Estimating exposure to mercury by different populations and establishing a safe level of consumption was another significant effort in the project (DEP 2007; DEP 2011; DEP 2013c). Results of the studies indicate that the vast majority of the man-made sources of mercury in Florida waters has global sources and that aquatic lakes and streams vary more because of their geochemistry than because of atmospheric loading. The TMDL report indicates significant reductions in mercury emissions have occurred in the last two decades.
No additional reductions will be required of local coal fired power plants due to recent large reductions arising from federal regulation (EPA 2013d) and the global nature of the sources in State waters. NPDES permit-holders will have no additional mercury limits imposed beyond currently enforced water quality criteria because of the limited impact of local atmospheric and point sources, and because of anticipated impending EPA regulations (EPA 2015b).
22.214.171.124. Current and Future: Mercury in LSJR Sediments
The influx of information about mercury sources and levels that will arise from the TMDL process will provide much needed information about the extent of the contamination throughout the state. In the LSJR, there is some mercury information but the amount of data is limited. For example, there is no information for the south mainstem, Area 4, for recent years and other areas in the LSJRB have limited numbers of samples. In addition, changes in standard methods of analysis make it difficult to track trends. The mercury database will be improved with the mercury TMDL process and future river status reports will summarize the results of that regulatory action.
Mean mercury concentrations in sediments collected from various sites along the LSJR Sites are given in Table 5.3. The distribution of mercury, the TEL, PEL, and hot spots in various years is shown in Figure 5.36. Mercury levels that exceed natural background levels and the most protective environmental guidelines are found throughout the mainstem. There are isolated locations in the LSJR, particularly in Rice Creek and the Cedar-Ortega system, where mercury occurs at concentrations high enough to impair the health of organisms. It is possible that mercury will bioaccumulate in those fish, crabs, and shellfish that spend most of their lives at these highly contaminated sites.
It should be noted that the toxicity pressure reflects the overall toxicological stress on the ecosystems of the river. It does not address human toxicity, which arises when we consume toxic metals that have found their way into the environment, via contaminated biota. Human health effects are discussed in the following section.
Because of the high degree of toxicity pressure due to mercury, the high numbers of sites that have mercury in sediments greater than background levels, and the high degree of potential human risk, the STATUS of mercury in sediments is unsatisfactory, and the TREND is unchanged.
Table 5.3 Average Mercury Concentrations and Percentage of Samples Exceeding Background and Sediment Quality Guidelines in the LSJR Sediments (see text in Section 5.2 for data sources)
|Average Conc., ppm||0.5||0.1||0.3||0.2||0.6||0.2||0.2||0.1||0.1||0.1|
|No. of Samples||13||28||143||52||214||40||45||28||25||16|
|% > TEL2||15%||32%||63%||75%||75%||53%||36%||39%||48%||38%|
|% > PEL2||15%||0%||6%||0%||30%||8%||2%||0%||0%||0%|
126.96.36.199. Mercury in LSJR Fish and Shellfish
The diverse types of fish that live in the LSJR were reviewed in Section 3 in this report. As noted, there is considerable overlap of freshwater, estuarine, and marine species in the dynamic LSJR system. In the following data sets, the marine and estuarine species associated with the LSJR were caught north of Doctors Lake. Of the marine and estuarine species discussed, King mackerel, Spanish mackerel, gag grouper, and bull shark are generally found offshore, while the others reside largely in coastal and estuarine waters. The freshwater species were caught south of Doctors Lake. The species that are reported are considered important because of their economic significance. Some species are also closely monitored because they are at high risk for elevated concentrations due to their large size and trophic status (Adams et al. 2003).
As shown in Figure 5.37, most species in the northern marine section of the LSJR, had low levels of mercury in their tissue, including blue crabs and oysters. The only data that exceeded FDOH’s most restrictive advisory levels for the general population were those reported in the Section 303(d) Impaired Waters listing for mercury. Those data, collected throughout Florida’s coastal and offshore waters, resulted in impaired designations for the marine and estuarine mainstem and seven tributaries north of Doctors Lake. The King mackerel and bull shark, top predator species that are large and long-lived, have significantly elevated levels compared to the other species. Levels in marine/estuarine species in the LSJR are comparable to or less than the averages for the individual species for the entire State of Florida (Adams et al. 2003). However, as discussed in Section 3, advisories have been issued for all Florida coastal waters for numerous species including Atlantic croaker, dolphin, gag grouper, King mackerel, sharks, red drum, southern flounder, spotted seatrout, and southern kingfish (FDOH 2016). Additional information about consumption advisories is available in Section 3 of this report.
In the fresh portions of the river south of Doctors Lake, the mainstem, tributaries, and large connected lakes, fish have been extensively sampled in the last 10 years (Figure 5.38). Levels exceeding the 0.3 mg/kg fish tissue criterion have been found primarily for largemouth bass, which caused the southern part of the LSJR mainstem, Lake Broward, and Crescent Lake to be designated as impaired. Not included in this discussion are several smaller, isolated southern lakes that have been listed as impaired due to elevated concentrations of mercury, again primarily in largemouth bass. As with the LSJR marine and estuarine fish, LSJR freshwater fish mercury levels are generally comparable to the rest of the state. Furthermore, the 1998-2005 national average for largemouth bass was 0.46 ppm, which is similar to LSJR values (Scudder et al. 2009).
There are a number of consumption advisories due to mercury contamination in fish in the LSJR region, and most fish contain at least small amounts of mercury. However, high levels of mercury in fish are found mostly in the top predators and in only a few of the fresh water bodies sampled. By consuming mostly lower-level predators and smaller, short-lived fish species (e.g., Atlantic croaker, flounder, sunfish) people can benefit from this healthy food source with minimal risk.
188.8.131.52. Point Sources of Mercury in the LSJR Region
In 2013, 558 pounds of atmospheric mercury emissions in the LSJR region were from four primary industries, including stone/clay/glass (30%), electric utilities (30%), primary metals (25%), and cement (15%). Emissions from gypsum and steel production have grown since 2008, offsetting reductions by the electric utility industry (Figure 5.39). St. Johns River Power Plant and Northside Generating Station reduced their mercury emissions by 71% between 2001 and 2013 (Figure 5.33). While 10 facilities reported mercury emissions, five were responsible for 99% of total atmospheric mercury emissions in 2013 (Figure 5.40).
Mercury releases into the LSJR and tributaries significantly dropped in 2004 with Seminole Generating station dramatically reducing its output of mercury. Coincident with reductions in atmospheric emissions since 2006, St. Johns River Power Park and Northside Generating Station steadily increased their discharges of mercury into surface water until 2011. However, in the subsequent two years there was a dramatic decrease in mercury discharges by that facility. Total discharges of mercury into the LSJR have been reduced by nearly 75% since 2001 (Figure 5.41). The RSEI model of chronic human health toxicity indicates that mercury releases to water by Seminole Electric is among the top potential risks compared to all releases in the region (EPA 2013e). However, we are unable to fully assess the importance of mercury because St. Johns River Power Park/Northside Generating Station, a major discharger, is not included in the RSEI modeling (see Section 5.3).