Assessing the ecological risk of heavy metal sediment contamination from Port Everglades Florida USA

Digestion and analysis

Subsamples of sediment were taken as previously described in White (2021). Specifically, subsamples of sediment (1 cm³, ∼2.0 g) at 5 cm intervals were taken along the entire length of each core from surface to base. Each sample was washed three times with ultrapure deionized water (18.2 M Ω) from a Barnstead water purification system. The sediments were pre-dried overnight in a VWR drying oven for 18 h at 80 °C, followed by 5 h in a Fischer Scientific isotemp vacuum oven (model 282A) at 80 °C at a pressure below 10⁻² torr, equipped with a Welch 1400 DuoSeal vacuum pump and a liquid nitrogen trap that resulted in completely dried samples. The dry weight of each sample was recorded and US Environmental Protection Agency (EPA) digestion method 3050B was used (Environmental Protection Agency (EPA), 1996).

Heavy metal analyses were performed as the University of Southern Mississippi’s Center for Trace Analysis as previously described in White (2021) by using a sector-field inductively coupled plasma mass spectrometer (ICP-MS) (ThermoFisher Element XR) with a Peltier-cooler spray chamber (PC-3; Elemental Scientific, Inc.). Prior to analysis, digested samples were diluted 5-fold in 0.64 M ultrapure nitric acid (Seastar Baseline) containing 2 ppb indium as an internal standard. Diluted samples were held in acid-washed Teflon autosampler vials. Mass spectrometer scans were performed in low (Cd-111, Hg-199,200,201,202, Pb-208), medium (Al-27, V-51, Cr-52, Mn-55, Fe-56, Co-59, Ni-60, Cu-63, Zn-66), and high (As-75, Se-77,82) resolution, depending on the isotope. Mo-98 was monitored to correct for MoO⁺ interference on Cd. Standardization was by use of external standards, with a high standard and a blank re-run every eight samples. For the elements (Hg, Se) where multiple isotopes were determined, no significant analytical differences were noted between the isotopes. Two USGS reference water concentrations were also assessed as part of each analytical run to verify the standardization. In several cases, sample calibration was also verified by standard additions. Blanks of ultrapure deionized water, hydrogen peroxide, and trace metal basis nitric acid (10%) were used for quality control purposes. Detection limits were calculated as three times the standard deviation of the blank (Table S1). Standard reference material SRM 2702-Inorganics in Marine Sediment (MilliporeSigma) was used to evaluate reliability of the analytical method and calculate percent recoveries (Table S2).

West Lake control site and continental crust composition

A control site for comparison with the port sites was established at West Lake (WL), located approximately 3 miles south of the port, outside of its influence. This site is part of West Lake Park, which includes a three-mile strip of mangrove estuary and uplands where no motorized boats are permitted. Since comprehensive background values for marine sediments for all 14 heavy metals have not been determined for this area, the port and reef heavy metal concentrations were also compared to the two cores that were collected from WL.

The elemental composition of the Earth’s upper crust, also referred to as background elemental composition concentrations, is used as a tool for assessing geochemical anomalies (heavy metal contamination) (Table S3). Since comprehensive background values for marine sediments for all 14 heavy metals have not been determined for this area, the concentrations were compared to continental crust values derived from the post-Archean Australian average shale (PAAS), European shale composite (ES), and North American shale composite (NASC) (Taylor & McLennan, 1995; Al-Mutairi & Yap, 2021). The sediment cores collected were at the coastal margin on continental crust. The transition between continental crust and oceanic crust is several miles from shore and any oceanic crust weathering is highly unlikely to impact the coastal sites (Webb, 0000; Britannica, 2023).

Geo-accumulation index

The geo-accumulation index (I_geo) is a quantitative measure of the degree of contamination in sediments (Förstner, 1980) and was used here to measure the pollution intensity at individual sample locations. The I_geo is calculated as follows: ${\displaystyle {\mathrm{I}}_{geo}=log2\left(\frac{{\mathrm{C}}_{\mathrm{n}}}{1.5x{B}_n}\right).}$

C_n = concentrations within the sediment cores and B_n = background continental crust levels. Rudnick & Gao (2014) provide values that quantify the degree of contamination: 4-5: strongly to extremely contaminated; 3-4: strongly contaminated; 2-3: moderately to strongly contaminated; 1-2: moderately contaminated; 0-1: uncontaminated to moderately contaminated; <0: uncontaminated.

Pollution load index

The pollution load index (PLI) was developed by Tomlinson et al. (1980) and is calculated using contamination factors (CF), represented by the concentrations of the sample metals and the background continental crust values C_metal/C_background. The calculation for PLI is: ${\displaystyle \mathrm{PLI}={\left({\mathrm{CF}}_1\mathrm{x}{\mathrm{CF}}_2\mathrm{x}{\mathrm{CF}}_3\mathrm{x}\dots {\mathrm{CF}}_{\mathrm{n}}\right)}^{1/\mathrm{n}}}$

where n = number of elements. This approach assesses total contamination load within the sediment and provides a PLI value that explains overall metal pollution within each sample. A sample with a PLI>1 is classified as polluted while a sample with a PLI<1 indicates no contamination (Tomlinson et al., 1980; Ray et al., 2006; Badr et al., 2009).

Threshold effect levels (TEL) and Probable effect levels (PEL)

The Florida Department of Environmental Protection created sediment quality assessment guidelines (SQAGs) to address coastal ecosystem contamination concerns. Numerical SQAGs were derived for nine metals (As, Cd, Cr, Cu, Hg, Ni, Pb, tributyltin, Zn) that occur in Florida coastal sediments. A threshold effect level (TEL) and probable effect level (PEL) were developed for these metals as powerful tools to assess contaminant levels in sediment (Macdonald et al., 1996). The TEL is the concentration below which adverse effects rarely occur to benthic organisms, while the PEL is the concentration above which adverse effects frequently occur (Geoenvironmental Engineering, 2015; Thompson & Wasserman, 2015).

Potential ecological risk index

The potential ecological risk index (PER) determines the degree of contamination for combined metal concentrations within each sediment sample (Guo, Xianbin & Zhanguang, 2010). The PER is calculated as: ${\displaystyle \mathrm{PER}=\sum _{}^{}\mathrm{E}}$ ${\displaystyle \mathrm{E}=\mathrm{TC}}$ ${\displaystyle \mathrm{C}={\mathrm{C}}_{\mathrm{a}}/{\mathrm{C}}_{\mathrm{b}}}$

where C_a = element content within sample, C_b = reference value of the element, and T = toxic response factor for metals: Mn and Zn =1, Cr =2, Cu and Pb =5, Ni =6, As =10, and Cd =30 (Hakanson, 1980; Fu et al., 2009; Guo, Xianbin & Zhanguang, 2010; Cao, Hong & Liu, 2015).

Ecological indices

Every ecological index has limitations as each assesses different heavy metal pollutants of importance. Geo-accumulation index (I_geo), potential ecological risk (PER), pollution load index (PLI), and enrichment factor (EF) are widely used for the assessment of the degree of pollution and health status of sediment. The advantage of the Igeo and EF indices is that they can identify whether the level of a specific heavy metal in sediment is due to anthropogenic or natural input or a combination of both (Schropp et al., 1990; Shafie et al., 2013; Zahra et al., 2014; Abdullah, Sah & Haris, 2020). The PLI identifies overall/total heavy metal contamination, and the PER identifies degrees of total metal contamination (Jahan & Strezov, 2018; Ali et al., 2022; Ranjani et al., 2021).

Concentrations of Fe and Al were not assessed in this study so EF values could not be determined. These two metals are most used for normalization to calculate the enrichment factor (EF) since normalization is used to differentiate if the metal source is of anthropogenic or natural origin.

Statistical analysis

Statistical analyses were performed in Microsoft Excel 365 (Microsoft 2018) and R Studio 2023.03.1+446. For all indices (geo-accumulation, pollution load, potential ecological risk) at all depths, a one-sided 95% T-confidence interval was used to determine the lower 95% confidence bound. Those values were compared to reference values (TEL, PEL) for a determination of the minimal potential contamination (to 95% confidence). Differences among replicate cores were determined using matched pairs t-tests with Bonferroni-Holms correction and significance level 0.05. For statistical purposes, half of the limit of detection was used for non-detected (nd) samples. Differences in median heavy metal concentrations at different depths at each site and for each contaminant were assessed using a Kruskal-Wallace test with Wilcoxon Rank-Sum and Bonferroni-Holms correction follow-up at significance level 0.05.

Heavy metal concentrations in port, control, and reef sites

All port sediment cores collected were from an area that has a significant disturbance signature due to past dredging. Concentrations of all 14 heavy metals were detected in all sediment cores and surface reef sediment (Tables S4–S17). Table 1 illustrates the geometric mean heavy metal concentrations, from greatest to least, across each sediment core while Table 2 shows all sediment concentration ranges, arithmetic means, and geometric means per location. Metal concentrations at respective depths across all cores per location revealed depth variability in most metals at only DCC and PHQ (Table S18). Due to inconsistencies of metal concentration variation by depth, data were compiled across all 10 port sediment cores and all subsamples. Likewise, the data were aggregated for all the subsamples of the two control site cores and the two reef locations. Port sediment cores have the highest geometric mean concentrations in eight of the 14 heavy metals (As, Cd, Cr, Co, Mo, Ni, Se, and V), while the control site has the highest concentrations for five heavy metals (Cu, Pb, Mn, Hg, and Zn), and the reef sediment sites have the highest for Sn. Mercury, Cd, Co, and Se were consistently the lowest concentrations across all locations—port, control, and reef. The greatest similarities in geometric mean concentrations among the three locations were found for As, Cd, Cr, and Hg.

Molybdenum geometric mean concentrations varied widely per location, as much as four orders of magnitude. Although Mo is moderately to highly represented at all the port sites (4.96 µg/g), it is the second lowest heavy metal concentration at the RF site (0.0021 µg/g), and the third lowest concentration at the control site (0.206 µg/g).

The highest geometric mean metal concentration across all sites is observed for Mn (45.4 µg/g) at the control site followed by a close grouping of V (10.8 µg/g), Zn (10.2 µg/g), As (8.82 µg/g), and Cr (6.88 µg/g) (Table 2).

Continental crust value comparison

Port sediment samples display the highest heavy metal concentrations above continental crust values (Table S3). Of the 258 port, 34 control, and six reef sediment samples analyzed, As concentrations exceed continental crust value (1.5 µg/g) 89%, 100%, and 100% of the time, respectively. Molybdenum concentrations exceed the continental crust value (1.5 µg/g) 79%, 6%, and 0%, at the three respective locations. Tin concentrations exceed crust value (5.5 µg/g) 22%, 3%, and 17% respectively, while Cu (25 µg/g) exceed 12%, 6%, and 17% at the respective locations (Tables S4–S17).

Geo-accumulation index (I_geo)

Geo-accumulation index measures the pollution intensity of individual sample locations and is a quantitative measure of the contamination degree in sediments relative to background continental crust values (Förstner, 1980). Arsenic is the only heavy metal that exhibits strongly to extremely contaminated geo-accumulation values (4–5) in the port and control sites but moderate to no contamination (0–2) at the reef sites. All sites have at least moderately contaminated geo-accumulation values of As. Statistical analysis indicates that the average As geo-accumulation values of the duplicate cores at the DCC (1.85–5.05) and STB (1.07–2.79) sites show contamination throughout the cores, while the PEC (0–2.58) and PHQ (0–4.02) cores mostly have As contamination, with some sections having no contamination. The control WL cores, and the two RF sites show moderate As contamination with values ranging 0.74–2.50 and 0.53–1.68, respectively (Table S19).

Molybdenum also exhibits strongly to extremely contaminated geo-accumulation values (4–5) in all port cores, but shows no contamination in the control and reef sites. All sites have at least moderately contaminated geo-accumulation values of Mo. Statistical analysis indicates that the average Mo geo-accumulation values of the duplicate cores at the STB (1.91–3.71) site show moderate to strong contamination throughout the cores while the DCC (−2.73–4.71), PEC (0–2.70), and PHQ (0–4.99) cores mostly have Mo contamination with some sections having with no contamination. The control WL cores shows no overall Mo contamination (Table S19).

Pollution load index (PLI)

The ratio of metal concentration and background continental crust value yields the pollution load index. Based on the PLI, the DCC and PEC port cores have polluted surface sediments. DCC and PEC sediment cores have a PLI>1.00 for DCC1 at 5 cm, DCC2 at 20 and 25 cm, PEC2, and PEC 3 at 5 cm. No significant PLI>1.00 were found for PHQ, STB, WL, and the RF sites, indicating no pollution (Table S20).

Threshold and probable effect levels (TEL and PEL)

Threshold effect level (TEL) and a probable effect level (PEL) have been derived for nine metals (As, Cd, Cr, Cu, Hg, Ni, Pb, tributyltin, Zn) that occur in Florida coastal sediments (Macdonald, 1994). Of the 258 port, 34 control, and six reef sediment samples collected, each site has varying TEL and PEL values for As. All four port sites have As concentrations above TEL (7.24 µg/g) and PEL (41.6 µg/g). Within the port, 96% of DCC samples are above As TEL, while 38% are above PEL. The PEC site has 38% of its samples above TEL and 0.9% above PEL. PHQ has 78% of its samples above TEL and 4% above PEL, while STB has 88% of its samples above As TEL and 4% above PEL. The control site, WL, has 58% of its samples above TEL and none above PEL. The reef sites, RF, have 17% above As TEL and none above PEL (Table S21).

Copper exceeds the TEL (18.7 µg/g) in samples at all port locations except one site (PHQ). The port cores have up to 60% of samples that exceed Cu TEL, while the control site has 11%, and the reef 17%. Two port sites, PEC and STB, have 2% and 8% of samples, respectively, with Cu above PEL (108 µg/g) but the control reef sites have none.

Port cores also have up to 22% of samples exceeding Ni TEL (15.9 µg/g) while 6% of the control samples exceed TEL. All port sites (up to 27% of samples) have samples exceeding Zn TEL (124 µg/g), while only the port DCC site has samples (2%) above PEL (271 µg/g). The Zn concentrations at the control site do not exceed TEL. Mercury concentrations exceeding TEL (0.13 µg/g) are present in 14% of samples at only one port site, PEC, and in 17% of samples of the control site, WL. Reef sites do not exceed TEL for Ni, Zn, and Hg. (Tables S4–S17).

Potential ecological risk (PER)

The combined metal concentrations within each sediment sample determines the potential ecological risk (Guo, Xianbin & Zhanguang, 2010). Sediment samples are ranked within five levels, ranging from low to moderate to considerable to high to significantly high. All port sites exhibit moderate to significantly high potential ecological risk. Statistical analyses show that the average PER values for DCC (76.5–755), PHQ (12.1–427), and STB (127–368) sites range from moderately to significantly high-risk levels; PHQ has two five cm sections (5 and 10 cm) with low PER (<40). PEC (11.1–441) has moderate to high risk levels throughout the cores and significantly high (332 for PEC 3 and 441 for PEC 2) at the surface (top 5 cm). The control site, WL, exhibits moderate to considerable potential ecological risk (28.2–137) while the reef sites display low to moderate risk 44.6 and 57.7 (Table S22).

Heavy metal concentrations among duplicate/triplicate cores

Port sample locations PEC, PHQ, and WL cores have no significant differences among their duplicate cores based on the geo-accumulation, pollution load, and pollution ecological risk indices. DCC cores 1, 2, and 3 show significant differences in the geo-accumulation index for Cd, while STB cores 1 and 2 show significant differences in the geo-accumulation and the ecological risk indices for Cu, Mo, Se, and Zn (Table S23). Except for the STB cores, there are little to no differences among the duplicate cores collected. Even so, it is still advisable to collect multiple cores per location for reproducibility.

Arsenic contamination in sediment

Arsenic is the contaminant that exceeds continental crust values in almost all the locations (89% port, 100% WL, and 100% RF of samples). The DCC locality has the highest concentrations of As where 96% of its samples are above TEL and 38% above PEL values at depths between 10–50 cm. The geo-accumulation index indicates strong to extreme As contamination in the port and WL, and moderate contamination at the RF. Results from port localities (PEC, PHQ, STB, and DCC) show 38–96% of the samples exceed the TEL value and 0.9–38% exceed the PEL value. The WL and RF localities exceed TEL values in 58% and 17% of their samples, respectively (Tables S4–S17).

Port sediments were expected to have higher concentrations of heavy metals due to 100 years of port activities. However, it was unexpected that the WL control site has As concentrations above TEL values, since it is distant from the port with no motorboat activity. Even so, port sediments have much higher overall As concentrations throughout their cores (Fig. 2) compared to WL (Fig. 3). WL core sediments have low heavy metal contamination except for As. The As concentrations in the port sediment (0.607–223 µg/g) are also among the highest determined in marine sediment found in ports and estuaries worldwide, except for a location in Estaque, France that ranged from 107–220 µg/g and at an English estuary that was contaminated with 1,740 µg/g with acid mine waste (Table S24).

In general, As is one of the most common toxins introduced anthropogenically through herbicides, pesticides, plant defoliants, wood preservatives, cattle-dipping vats, chicken litter, coal burning, trash incineration, water treatment sludge, and industrial sources (Macdonald et al., 1996; NRC, 2000; Rosen & Liu, 2009; Missimer et al., 2018). The As found in both the port and WL locations could be the result of runoff and atmospheric deposition. Between 1992–1993, As concentrations (13–51 µg/g) were detected in reclaimed water at Bonaventure Golf Course, located due west of Port Everglades along the L-35 canal, which flows to the Intracoastal Waterway and Port Everglades (Fig. 4). The As originates from the monosodium methanearsonate that was applied to the turf as an herbicide to control grasses and weeds (Swancar, 1996). Arsenic can also be released during municipal solid waste incineration and deposited back to the environment (Hu et al., 2015). During the high temperature incineration process, arsenic is predominantly present as As₂O₃(g) vapor and as a result, concentrations are usually elevated in local marine environments and bioaccumulate through trophic webs (Volesky, 1990; Hu et al., 2015).

One of the major sinks for As from municipal waste incineration is in sediments (Pierce & Moore, 1982). Broward County’s Environmental Monitoring Laboratory (BCEML) reported concentrations of As (0.623–19.4 µg/g) in marine surface sediment in the C-11, C-12, C-13 canals, and Intracoastal Waterway from 2005–2007 (Bernhard, 2014). In 2005, Bernhard (2014) reported As concentrations below 2.43 µg/g in surface marine sediment from multiple Intracoastal Waterway locations north of Port Everglades to the middle branch of New River. The results from this project indicate As concentrations that are up to ten times higher at certain depths in the port sediment (0.607–223 µg/g) while the mean value of 16.1 µg/g is in line with the higher concentrations of As determined compared to the BCEML data. The control site had approximately equal As concentrations (1.7–21.7 µg/g) and mean of 8.5 µg/g, while the reef site As concentrations were lower (2.41−8.55 µg/g) with a mean of 5.23 µg/g compared to the BCEML data (Table 2). The port and control site sediment cores were taken at variable depths, while the BCEML data were for surface sediment only, which could explain the differences in concentrations. This study indicates the importance of collecting sediment cores from the port since concentrations vary per depth and dredging events take sediment from variable depths.

In its inorganic forms, As is lethal to organisms and harmful to the environment (Jaishankar et al., 2014). Its toxicity to marine organisms is complex due to the existence of two inorganic arsenic species As (III), as arsenite, and As (V), as arsenate; both can be found in aquatic ecosystems (Liber, Doig & White-Sobey, 2011). Arsenite is more lipid soluble than arsenate and has the highest acute toxicity among the different As forms (Saha et al., 1999). Studies with freshwater invertebrates determined that arsenite is generally more toxic than arsenate (Borgmann, 1980; Spehar et al., 1980; Golding, Timperley & Evans, 1997). Various sublethal effects on behavior, growth, locomotion, reproduction, and respiration indicate that As is acutely toxic to marine organisms (Macdonald et al., 1996).

Molybdenum contamination in sediment

Molybdenum concentrations exceed continental crust values in 79% of all port sediment samples, only 6% in WL, and none in RF. Its concentrations are higher in the port sediment (Fig. 5) than WL (Fig. 6) and RF by as much as four orders of magnitude. The geo-accumulation index indicates strong to extreme Mo contamination in port sediment, low to moderate in WL, and none in RF. The Mo concentrations in the port sediment (nd-385 µg/g) are amongst the highest determined in marine sediment found in ports and estuaries worldwide, where values ranged 0.5–40 µg/g (Table S24).

In higher oxic conditions, Mo accumulates at concentrations close to the crustal concentration abundance (1.5 µg/g), while in anoxic and sulfidic conditions Mo is enriched (Crusius et al., 1996). The differences observed among the three locations for Mo may result from more oxic conditions found at the RF sites. Molybdenum concentrations in sediment, along with Fe distribution, is used to identify oxic conditions in marine systems (Scholz et al., 2017). Molybdenum can be transferred from the water column to sediments through particulate shuttle effects involving Fe and Mn and has been found in areas bearing manganese oxides (Crusius et al., 1996; Algeo & Lyons, 2006; Algeo & Tribovillard, 2009). To our knowledge, Mo concentrations and its possible effects have not been reported for Broward County, Florida.

Manganese contamination in sediment

Manganese has the overall highest geometric mean concentration (45.4 µg/g) of all the heavy metals analyzed (Table 2). Although Mn occurs naturally in sediments and soils due to the weathering of parent material, anthropogenic processes such as mining, smelting and addition of biosolids/organic wastes to agricultural areas increase Mn concentrations (Paschke, Valdecantos & Redente, 2005; Boudissa et al., 2006; Li et al., 2013). Concentration ranges of heavy metals in surface waters at six sampling sites near Port Everglades between 1975–1978 indicated Mn concentrations of nd-120 µg/L (Sonntag, 1980). In addition, in 1983 heavy metal concentrations of lithophilic samples at varying depths at the Dixie water-treatment plant site adjacent to US Highway 441, Broward County, were <1–26 µg/kg (Howie & Waller, 1986). In comparison, the sediment results from the port, control, and reef sites in this study indicate much higher Mn concentrations. The port samples have 1.61–204 µg/g (mean: 33.0 µg/g), the control 9.38–93.8 µg/g (mean: 52.0 µg/g), and the reef sites 10.1–24.1 µg/g (mean: 15.3 µg/g) (Table 2). These higher concentrations could be due to Mn sediment accumulation over time through runoff.

There is little data on how Mn affects the biota in marine environments and TEL and PEL values have not been established for Mn; however, tropical species appear to be more sensitive to Mn concentrations than temperate ones (Anastasi & Wilson, 2010). Even though Mn concentrations did not exceed the background continental crust value of 600 µg/g, the overall concentrations can be of concern to benthic organisms since studies indicate sensitivity of benthic organisms to certain Mn concentrations. Summer, Reichelt-Brushett & Howe (2019) showed reduced hard coral fertilization success after exposure to 164–237 µg/g of Mn, larval death at 7 mg/L, and adult coral tissue death at 0.7 mg/L (where mg/L is equivalent to µg/g). In our study, the concentration of Mn in the port sediment is 1.61–204 µg/g, 9.38–93.8 µg/g at the control site, and 10.1–24.1 µg/g at the reef. The proximity of the coral reef tract to the port warrants concern should Mn in the port sediment remobilize during dredging. All port and control site cores were collected along mangrove channels, and Mn has been associated with mangrove forest rhizospheres and high permeability of sediments (Guieros et al., 2003). Studies have shown that seagrass uptakes heavy metals (Smith et al., 2019) and mangroves contribute to heavy metal retention (Nguyen, Le & Richter, 2020). In the marine environment, aquatic plants play a crucial role in the uptake, storage, and recycling of heavy metals (Chandra & Kulshreshtha, 2004; Smith et al., 2019). The Mn concentrations in the port sediment (1.61–204 µg/g) are equivalent or even slightly below marine sediment found in ports and estuaries worldwide (0.4–4643 µg/g) (Table S24).

Copper contamination in sediment

Copper concentrations exceed continental crust values in 12% of all port sediment samples, 6% in WL, and 17% in RF samples. As many as 60% of port sediment samples exceed TEL values, and 11% of WL samples and 17% of RF samples exceed TEL values (Tables S4–S17).

Copper is a common metal that exists in three forms, Cu, Cu (I), and Cu (II). Its ionic forms are biologically available since it is essential for the proper growth of animals (Flemming & Trevors, 1989; Ikemoto et al., 2004). Weathering of copper-bearing metals, copper sulfides, and copper are natural sources of this element within marine ecosystems (Macdonald et al., 1996). Anthropogenic sources of copper include copper wire mills, coal burning, smelting, refining, and iron and steel producing industries (Health Canada, 2019). Copper is an essential micronutrient and is therefore easily and readily accumulated by marine organisms and plants (Macdonald et al., 1996). However, Cu can be toxic to bivalve mollusks, altering the biochemical and physical properties of the surface epithelium and disrupting membrane permeability (Cheng, 1979). Dean, Shimmield & Black (2007) discovered that a significant amount of Cu in marine sediment at fish farms originated from anti-fouling agents. These can be released in either solid or particulate form originating from traps, coated nets, or hard structures (i.e., container ships) into the marine environment (Claisse & Alzieu, 1993; Miller, 1994; Brooks, 2000; Morrisey et al., 2000; Solberg, Sæthre & Julshamn, 2002; Brooks & Mahnken, 2003).

Concentration ranges of Cu in surface waters at six sampling sites near Port Everglades between 1975–1978 were nd-55 µg/L (Sonntag, 1980). In the 1980s Cu concentrations in stormwater runoff from residential, commercial, and roadway lands in Broward County, Florida were nd-500 µg/L (Whalen & Cullum, 1988). Copper concentrations in highway runoff, shallow groundwater, and lithophilic samples in 1983, at varying depths at the Dixie water-treatment plant site adjacent to US Highway 441, Broward County, had values ranging from <10–100 µg/L, <10–40 µg/L, and <1–10 µg/kg, respectively (Howie & Waller, 1986). BCEML reported the concentration of Cu (3.3–413 µg/g) in marine surface sediment in the C-11, C-12, C-13 canals, and intracoastal waterway from 2005–2007 (Bernhard, 2014). In 2005, Bernhard (2014) reported Cu concentrations (6.05–153 µg/g) in surface marine sediment from multiple Intracoastal Waterway locations north of Port Everglades to the middle branch of New River. The results from this study indicate Cu accumulation in the port, control, and reef sites sediment similar to the 2005–2007 and 2014 studies, with the port values ranging from 0.004–215 µg/g (mean: 11.9 µg/g), the control from 0.260–30.4 µg/g (mean: 7.44 µg/g), and the reef sites from 0.510–28.6 µg/g (mean: 5.35 µg/g) (Table 2). The Cu concentrations in the port sediment are also equivalent or slightly below marine sediment found in ports and estuaries worldwide (0.06–1195 µg/g), except for a location at an English estuary that was contaminated with 2,398 µg/g from acid mine waste (Table S24).

Ecological impacts to benthic communities

The port sediment heavy metal concentrations have the potential to ecologically impact benthic organisms. Port Everglades, a commercial port, is intermittently dredged to accommodate the draft of large ships. The dredging of this port can present a contaminant threat to benthic organisms and the adjacent coral reef community, as most sediment collected in the port displays evidence of high heavy metal contamination and ecological risk.

The resuspension of contaminated dredged marine sediment and the remobilization of heavy metals can have adverse ecological effects on resident aquatic organisms. Wei et al. (2021) determined that dredged materials at the dumping area of Huangmao Island, China had a negative impact on the benthic community and that larger dumping amounts lead to higher heavy metal pollution risks. Kalnejais, Martin & Bothner (2010) determined that sediment resuspension was responsible for transferring significant amounts of metals (Fe, Mn, Ag, Cu, and Pb) to the mobile dissolved phase in Boston Harbor, Massachusett. Fetters et al. (2016) revealed that redeposited suspended sediment increased the bioavailability and toxicity of Zn, Cu, Cd, Pb, Ni, and Cr to marine Hyalella azteca (amphipod crustacean). Banks et al. (2012) investigated the effect of trace metals (including As) during an extended oxygen depletion event across the sediment-water interface in sediments from a highly metal contaminated estuary in southeast Tasmania, Australia. Where hypoxia developed, Mn and Fe were released from sediments at all sites, while As, Cd, Cu and Zn release was comparatively low. Ferrans et al. (2021) assessed the metal pollution risk during present and future dredging at Malmfjärden Bay, Sweden. Contamination and ecological risk factors indicated low Cr, Ni and Fe pollution concern and a medium risk for Pb and Zn during dredging activities.

More research could help to understand how dredging and other types of disturbances on marine geochemistry affect the resuspension and release/remobilization of sediment contaminants, such as heavy metals. Even so, it is known that the disposal of dredged material is the largest mass input of waste deposited into the oceans (Vivian & Murray, 2009).

The use of ecological indices is a common and accepted practice in assessing sediment heavy metal contamination and risks to the respective environments. However there are limitations using ecological indices such as geo-accumulation, pollution load, and potential ecological risk. These indices use sediment metal concentration data and synthesize it in an understandable way, but it can also be overly generalized. A single index alone does not provide the whole story of ecological risks and/or contamination. An unusually high reading at one location or depth could skew the index outcome. This is why multiple ecological indices are used to provide a more comprehensive assessment of contamination and help translate the heavy metal data into understandable measures of ecological risks.