T.F. Thigpen, A.Y. Oguma, and P.L. Klerks 22001144 EASETaEstRerNn BBIiOolLoOgiGstIST No. 2:N1–o1. 12 The Effects of Phenanthrene on the Benthic Macroinvertebrate Community of a Louisiana Swamp Tyler F. Thigpen1, Andrew Y. Oguma1,2,*, and Paul L. Klerks1 Abstract - This study explored the effects of phenanthrene on the benthic macroinvertebrate community using intact, field-collected microcosms from a site in the Atchafalaya River Basin (ARB), an expansive forested swamp in south-central Louisiana. Understanding the effects of polycyclic aromatic hydrocarbons on naturally occurring invertebrate communities in this ecosystem is important because the ARB is used for extensive oil and natural gas exploration and has experienced oil spills and leaks. In fall 2010, we collected intact sediment cores from a backwater site in the ARB and spiked the overlying water with nominal phenanthrene concentrations of 0 μg/L, 50 μg/L, or 100 μg/L for a 20-day static exposure. At the end of the exposure period, microcosms from the ARB with phenanthrene levels of 50 μg/L or higher had significantly fewer benthic macroinvertebrates than controls (F2,26 = 25.67, P < 0.0001). Our results also indicate that our methodology employing in situ-collected intact sediment cores with their naturally occurring benthic communities to use as laboratory microcosms may be a good technique for use in toxicity tests on naturally occurring benthic communities. Introduction The objective of our study was to assess the effects of phenanthrene (C14H10), a polycyclic aromatic hydrocarbon (PAH), on benthic macroinvertebrate communities collected from the Atchafalaya River Basin (ARB), Louisiana. We conducted this assessment with intact sediment cores as laboratory microcosms. Assessing the effects of a toxin at the community level on a relatively natural benthic community had not yet been conducted in the ARB. The use of microcosms in bioassays provides an efficient method for determining the effects of a contaminant on target communities rather than individual species (Pontasch et al. 1989). Balthis et al. (2010) observed that the effects of contaminated sediments on benthic invertebrates may be greater at the community level than at the individual level. Assessing community-level effects of contaminants on macrobenthos may provide information regarding the effects of toxins on ecosystem processes—including primary production, nutrient cycling, and energy flow—beyond the traditional, single species bioassays (Clements and Rohr 2009). The ARB (30º30'04.37"N, 91º43'52.09"W), located in south-central Louisiana (Fig. 1), is the largest contiguous bottomland hardwood forest in North America— more than 5100 km2 in size—and provides water for drinking, agriculture, and recreation (Bergstrom et al. 2004). The vast wetland ecosystem is comprised 1Department of Biology, University of Louisiana at Lafayette, Lafayette, LA 70504. 2Department of Biology, Western Connecticut State University, 181 White Street, Danbury, CT 06810. *Corresponding author - ogumaa@wcsu.edu. T.F. Thigpen, A.Y. Oguma, and P.L. Klerks 2014 Eastern Biologist No. 2 2 of diverse freshwater habitats ranging from hypoxic to oxygen-rich, including seasonally flooded swamps, and backwater and shallow headwater lakes interspersed with a wide-ranging network of dredged channels for oil and gas transport (Hupp et al. 2008). The ARB is also used for onshore oil and natural gas production (Demcheck and Swarzenski 2003, Hupp et al. 2008) and contains more than 300 active oil and natural gas wells (Holland et al. 1983) . Phenanthrene, an Environmental Protection Agency (EPA) priority pollutant, is a component of many PAH mixtures (Irwin et al. 1997, Shea et al. 2001, USEPA 1982). The last major spill documented in the ARB occurred at Myette Point as a result of a well blowout that released oil for five days in December 1996. An estimated 747,240 liters of gas condensate spilled into the Myette Point canal and adjacent forests. Penland et al. (1999) reported that this spill resulted in 43 ha of light oil-cover (only oil film present on water), 12 ha of moderate oil-cover (<50% of area covered in oil), and 12 ha of heavy oil-cover (>50% of area covered in oil). A survey conducted in winter 1997 documented plant mortality at 10%, 50%, and 90% in the light, moderate, and heavy oil areas, respectively (Penland et al. 1999). The Myette Point oil spill is an example of a large spill occurring within the ARB; however, benthic assessment and effects on benthic species were not conducted in association with the Myette Point oil spill. Given the large area potentially affected by spills, it is important to understand the impact of oil on the ARB. Studies of the effects of PAHs on bottomland hardwood forest communities are especially important, because oil persists longer in low energy ecosystems like marshes and bottomlands than in higher energy systems like lakes and rivers (Baca et al. 1985) thus increasing the likelihood of negative impacts. Findings from an evaluation of the concentrations of PAHs in sediment in the Atchafalaya National Wildlife Refuge, located in the ARB, showed that PAH levels at some sites exceeded levels that, according to sediment-quality guidelines, are often associated with toxic effects (Shea et al. 2001). Shea et al. (2001) classified PAH levels at sites near an inactive oil pit, a barge, and an active oil platform as extremely high and lethal to wildlife. However, they did not address the potential for deleterious effects on invertebrates. Phenanthrene is known to have adverse effects on macroinvertebrates, including Hyalella azteca Saussure (Lee at al. 2002, Lotufo and Landrum 2002) and Procambarus clarkia (Girard) (Red Swamp Crayfish) (Umejuru 2007), which occur in the ARB (Sklar 1985). Toxicity of phenanthrene in sediment has also been reported for species of oligochaetes (Phylum Annelida) and chironomids (Subclass Insecta, Order Diptera, midges), the two taxonomic groups most common in fine, organic-rich sediments such as those occurring in the ARB (Lotufo and Fleeger 1996, Verrhiest et al. 2001). However, it is difficult to extrapolate from studies of single species in formulated sediments or modified natural sediments to community-wide exposure to phenthrene in natural sediment because the organic content of sediment is known to affect PAH bioavailability and bioaccumulation (Lamy-Enrici et al. 2003, Mitra et al. 2000). Because benthic species differ in 3 T.F. Thigpen, A.Y. Oguma, and P.L. Klerks 2014 Eastern Biologist No. 2 their sensitivity to PAHs (Triffault-Bouchet, 2005), bioassays using a few standard species cannot provide an accurate impact assessment at the community level. Effects on benthic communities are apparent at much lower levels of PAH-exposure than the threshold levels reported for effects on individual species in standard sediment-bioassays (Balthis et al. 2010). The legacy of oil exploration in the ARB necessitates a clear understanding of the effects of phenanthrene on communities of benthic invertebrates. We collected sediment cores from the ARB during October 2010 to examine the effect of phenanthrene on the benthic invertebrate community. To our knowledge, no previous study had investigated the effects of PAHs on benthic communities occurring in the ARB, a group of organisms responsible for 15% of the total community respiration (Sklar 1985). A secondary goal of this study was to explore a technique in which we collected in situ sediment cores and their associated benthic communities and used them directly as laboratory microcosms for a toxicity bioassay. In many microcosm studies, the sediment used is either natural sediment that has been manipulated by sieving, homogenizing and freezing, and/or fortified with food (Bhattacharyya et al. 2003, Caliman et al. 2007, Clément et al. 2004), or artificially formulated sediment obtained from a laboratory supply company (Clément and Cadier 1998). Day et al. (1995) observed that the manipulation of sediment prior to its use in sediment toxicity tests is known to affect toxicity outcomes. In addition to using disrupted or artificial sediment, toxicity bioassays vary in the selection of study organisms used from species collected from the wild (Caliman et al. 2007, Galar-Martínez et al. 2008) to artificially assembled communities of laboratory-cultured species (Clément and Cadier 1998, Clément et al. 2004, 2005). The presence of native organisms in sediment used in bioassays affects the results of sediment-toxicity tests (Day et al. 1995, Reynoldson et al. 1995). To our knowledge, ours is the first study for the ARB to use undisturbed, field-collected sediment and the associated benthic community. Our technique may provide important insights into the effects of toxins on ecologically relevant communities of benthic invertebrates and demonstrate the utility of the method for laboratory testing. Materials & Methods Sediment collection We collected 35 sediment cores in October 2010 from a 0.2-ha backwater site in the ARB (Fig. 1). We chose the site on the basis of a preliminary in situ assessment of benthic macroinvertebrate abundance and diversity in the area. We used a sediment-coring tube 15 cm in length with a 5.1–cm internal diameter, and a 6.1–cm external diameter. We pushed the tube 8 cm into the sediment to collect a total volume of approximately 162 cm3. We capped the tube top with a 6.4–cm diameter PVC cap (PN 447-020HC; Mueller Streamline, Memphis, TN) to create the suction necessary to remove the core from the surrounding sediment. While removing the tube from the surrounding sediment, we slid a piece of plastic under the bottom of the core. After retrieving the core from the sediment, we capped the bottom T.F. Thigpen, A.Y. Oguma, and P.L. Klerks 2014 Eastern Biologist No. 2 4 of the tube with a 6.4-cm diameter PVC cap. The resulting cores contained approximately an 8-cm column of sediment and 6 cm of overlying ambient water. We removed the top cap and covered each core with Parafilm® to conserve the ambient water while transporting cores back to the laboratory. Community assessment To assess the benthic macroinvertebrates present in the cores, we processed 5 cores approximately 24 h after collection. We wetsieved sediment with a 500-μm-mesh sieve, sorted benthic organisms, and preserved them in 10% neutral buffered formalin for 24 h. Organisms were then transferred to 50% ethyl alcohol for storage and later identification. We identified benthic invertebrates to order or family using Merritt and Cummins (1996), and Thorp and Covich (2010). Exposure We incubated collected cores for 7 d at 20.5 °C, the air temperature in the field when cores were collected, to allow them to equilibrate. Following the incubation period, we employed treatment interspersion (Hurlbert 1984); we randomly assigned each core (microcosm) to one of four treatments—control, acetone control, 50 μg/L of phenanthrene (phen), and 100 μg phen/L—and applied the treatments to the overlying, ambient water. To determine the appropriate amount of phenanthrene to add, we collected 6 cores Figure 1. Maps showing sediment collection site (black circle) in the Atchafalaya River Basin (top) and its location in the Mississippi River watershed (bottom). 5 T.F. Thigpen, A.Y. Oguma, and P.L. Klerks 2014 Eastern Biologist No. 2 from our original collection site and measured the volume of overlying water (mean = 0.4864 ± 0.0327 L). We made a stock solution of phenanthrene by dissolving solid phenanthrene (98% purity; Sigma-Aldrich, St. Louis, MO) in acetone to a nominal concentration of 785.5 μM. We added this phenanthrene stock solution directly to the overlying water of microcosms (1.7 μL for 50 μg phen/L-treatment cores, n = 10 and 3.4 μL for 100 μg phen/L-treatment cores, n = 10). Nothing was added to control treatments (n = 5). We added 3.4 μL of acetone (the amount used in the 100 μg phen/L-treatment) to the overlying water for a solvent control-treatment (n = 5). We used only 5 cores for each of the controls because we had limited supplies and space to transport the microcosms. We chose to spike microcosm overlying water to nominal phenanthrene concentrations of 50 μg/L and 100 μg/L because these levels are comparable to those that have been found in oil spills, service station run-off, and roadside streams (Irwin et al. 1997, Lefcort et al. 1997, Scoggins et al. 2007). We covered the 30 microcosms with 2 layers of 500 μm-mesh plastic screen which we offset to reduce the mesh size and prevent emerging invertebrates from escaping the tubes and affixed them to the microcosms with rubber bands. We placed the cores back in the incubator at 20.5 °C on a 12-hour light/ 12-hour dark setting for the 20-day exposure period. We used a static system for exposure in order to mimic conditions at our collection site—a hypoxic backwater swamp ecosystem. During the 20-day exposure period, we counted and recorded emerging macroinvertebrates every 2–3 days. All emerging invertebrates were adult chironomid midges which we easily quantified by counting them through the mesh screens. We removed emerged individuals to ensure they were counted only once. At the end of the 20-day exposure period, we wet-sieved each microcosm with a 500-μm sieve, and collected and identified invertebrates to order or family using Merritt and Cummins (1996), and Thorp and Covich (2010). Statistical analysis We conducted all statistical analyses using Statistical Analysis Software (Version 4.2; SAS Institute, Inc., Cary, NC) with an alpha level of 0.05. We used a one-way analysis of variance (ANOVA) model to test for differences among treatments in the numbers of macroinvertebrates remaining following exposure and to assess numbers of insects emerging during exposure. In both models, the independent variable was treatment (control, acetone control, 50 μg phen/L, and 100 μg phen/L). Tukey HSD-adjusted post-hoc tests were used to identify differences among treatments. We used additional one-way ANOVAs to test for differences in total macroinvertebrate densities among the initial community assessment and both control treatments (control and acetone control). We examined diagnostic outputs to assure that all data analyzed with parametric ANOVAs satisfied assumptions of homogeneity of variance and normal distribution. Results Our collected cores contained a substantial number of macroinvertebrates (mean = 157 individ/cm3 sediment). The majority (79%) of individuals were T.F. Thigpen, A.Y. Oguma, and P.L. Klerks 2014 Eastern Biologist No. 2 6 Initial n = 5 Control n = 5 Acetone control n = 5 50 μg phen/L n = 10 100 μg phen/L n = 10 Chironomidae 122 ± 5.2 2 ± 1.2 22 ± 7.5 53 ± 53 0 ± 0 Tubificidae 27 ± 5.4 17 ± 7.8 16 ± 6.4 7.5 ± 2.1 0 ± 0 Ceratopogonidae 4 ± 2.9 5 ± 3.1 0 ± 0 0 ± 0 0 ± 0 Hemiptera 3 ± 2 11 ± 11 0 ± 0 0 ± 0 0 ± 0 Asellidae 1 ± 1 0 ± 0 0 ± 0 0 ± 0 0 ± 0 Mollusca 0 ± 0 0 ± 0 1 ± 1 0 ± 0 0 ± 0 Total 157 ± 4.4 35 ± 7.9 39 ± 5.6 8 ± 2.3 0 ± 0 Non-dipterans 31 ± 4.3 28 ± 9.4 17 ± 6.4 7.5 ± 2.1 0 ± 0 Table 1. Macroinvertebrate densities in cores immediately after collection (initial) and at the end of the phenanthrene (phen) exposure experiment. Cores were left untreated (control), received phenanthrene to a concentration of 50 μg phen/L, received phenanthrene to a concentration of 100 μg phen/L, or received only the acetone used as solvent for the phenanthrene treatments (acetone control). Values are mean ± SE; n = the number of cores per treatment. Non-dipterans = the sum of Tubificidae, Mollusca, Asellidae, and Hemiptera. Figure 2. After 7 days of incubation plus 20 days of treatment exposure, we sieved microcosms , and living organisms were counted. The top graph shows the number of organisms present in sediment for experimental treatme n t s : co n t r o l , acetone control, phenanthrene (phen) treatments 50 μg phen/L, and 100 μg phen/L. The bottom graph shows the number of emerging invertebrates (metamorphosing individuals) for experimental treatments. Values are means and error bars represent standard error. Letters (A, B) denote significant differences in treatments. 7 T.F. Thigpen, A.Y. Oguma, and P.L. Klerks 2014 Eastern Biologist No. 2 Chironomidae (chironomids) with a substantial number (18%) of Tubificidae (tubificids), and a small number (3%) of both Asellidae (isopods) and Ceratopogonidae (no-see-ums or biting midges) (Table 1). After the 7-day equilibration plus 20-d treatment period, benthic macroinvertebrate densities were significantly lower in both sets of control groups (control and acetone control) relative to initial densities (Table 1; F2,12 = 167.68; P < 0.0001; a priori contrast of initial vs. control and acetone control, F1,12 = 331.63; P < 0.0001). There was no significant difference in total macrobenthos density between the two control groups (post-hoc pairwise comparison using Tukey HSD: control = acetone-control > 50 μg phen/L = 100 μg phen/L). At the end of the 20-d treatment period, the two phenanthrene treatment groups had significantly fewer benthic macroinvertebrates than did cores in the two control treatments (F2,26 = 25.67; P < 0.0001). Macroinvertebrates were completely absent from the 100-μg phen/L-treatment cores. Significantly (90%) fewe r adult chironomids and no-see-ums emerged in the phenanthrene treatments than in the acetonecontrol treatment (F2,26 = 10.48; P = 0.0001; post-hoc pairwise comparison using Tukey HSD: control = acetone control > 50 μg phen/L = 100 μg phen /L) (Fig. 2). Discussion Our findings suggest that exposure to phenanthrene at concentrations as low as 50 μg phen/L can cause a substantial decrease in the number of native macrobenthic organisms alive at the end of the 7-day equilibration plus 20-day exposure period. We also found a significant reduction in the number of adult chironomids and no-see-ums emerging from the phenanthrene-dosed cores. To our knowledge, this is the first study of phenanthrene toxicity on naturally occurring benthic macroinvertebrate communities. This study was conducted with ecologically relevant nominal treatment levels of phenanthrene: 50 and 100 μg/L in overlying water that mimic the mean concentration of 52.65 μg/L recorded in sediment and tarballs in association with the MC 252/Deepwater Horizon oil spill in Louisiana, Florida, and Alabama (Rosenbauer et al. 2010). Additionally, Leftcort et al. (1997) recorded oil concentrations of approxi mately 100 μg/L in service station run-off. Phenanthrene-toxicity has been assessed for individual species of benthic invertebrates, and assessments at the community level have been conducted at sites contaminated with complex PAH mixtures (i.e., not limited to phenanthrene). The results of these studies indicate that chironomids tend to be more sensitive than oligochaetes to phenanthrene or PAH mixtures (Bhattacharyya et al. 2003, Scoggins et al. 2007). This difference might explain why the initial and control cores in our study had a higher ratio of chironomids to oligochaetes (2:1) than the 50-μg-phen/L cores (1:15). Our methodology used intact cores of field-collected sediments as laboratory microcosms for toxicity testing. Clements and Newman (2002) criticized microcosm research and suggested that the decreased variability and increased reproducibility comes at the expense of ecological relevance. However, our technique maintains T.F. Thigpen, A.Y. Oguma, and P.L. Klerks 2014 Eastern Biologist No. 2 8 ecological relevance because it tests a more-or-less natural community that occurs within the microcosms. Our results suggest that the methodology may be useful for toxicity testing in other aquatic ecosystems. Our control cores, sieved after 27 days in the laboratory, contained significantly fewer organisms than our initial cores which we sieved about 24 h after collection in the field. Dipteran emergence (whose larvae reside in the sediment) or mortality might explain the large drop in invertebrate numbers between the initial and control cores, because differences in benthos density between initial and control cores were no longer evident when the dipteran groups (chironomids and no-see-ums) were excluded from our analysis. Because we did not monitor adult insect emergence during the 7-day equilibration period we cannot determine whether the dramatic change in community structure was due to chironomid emergence or mortality. For future studies using this method, we recommend monitoring insect emergence during the initial acclimation period in addition to the period of toxicity testing. The sediment cores we used to establish our microcosms were relatively small, with a diameter of about 5 cm. Henke and Batzer (2005) compared different methods for sampling macroinvertebrates in a wetland in the Georgia Piedmont region, and found that small corers capture relatively few individuals and tend to miss rare or mobile organisms. Our small cores contained a substantial number of individual macroinvertebrates (30–34 individuals and an estimated 5–8 species per core in the five cores processed shortly after collection), indicating that such small cores can work well for freshwater swamps with a rich macrobenthic community. Sklar’s (1985) findings indicated that the use of small corers is suitable for sampling the abundant and diverse invertebrate assemblages in a Louisiana bottomland hardwood swamp. In conclusion, our results indicated that phenanthrene is detrimental to naturally occurring macroinvertebrate communities in a forested swamp in south-central Louisiana. Conducting further research to determine the effects of PAHs on invertebrates in the ARB ecosystem and on benthic communities in general is imperative. We recommend our method for conducting benthic macroinvertebrate microcosm experiments on naturally occurring communities. Acknowledgments John W. McCoy at the US Geological Survey, National Wetlands Research Center (USGS NWRC) provided equipment. Jeromi Heffner provided laboratory support for processing microcosms. Kate Spear provided writing support. Darren Johnson, of 5 Rivers Services at the USGS NWRC, provided statistical support. Brad Glorioso of USGS NWRC provided GIS support. The University of Louisiana at Lafayette (UL Lafayette) Graduate Student Organization provided research funds for this project and UL Lafayette Center for Ecology and Environmental Technology provided the facilities necessary to conduct this experiment. 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