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Immunomodulatory effects of exposure to polychlorinated biphenyls and perfluoroalkyl acids in East Greenland ringed seals (Pusa hispida)
Environmental Research 151 (2016) 244–250
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Immunomodulatory effects of exposure to polychlorinated biphenyls and perfluoroalkyl acids in East Greenland ringed seals (Pusa hispida) Milton Levin a,n, Erika Gebhard a, Lindsay Jasperse a, Jean-Pierre Desforges b, Rune Dietz b, Christian Sonne b, Igor Eulaers b, Adrian Covaci c, Rossana Bossi d, Sylvain De Guise a a
Department of Pathobiology and Veterinary Science, University of Connecticut, 61 North Eagleville Road, Storrs, CT 06269-3089, United States Department of Bioscience, Arctic Research Centre (ARC), Aarhus University, Frederiksborgvej 399, PO Box 358, DK-4000 Roskilde, Denmark c Toxicological Centre, Department of Pharmaceutical Sciences, University of Antwerp, Universiteitsplein 1, BE-2610 Wilrijk, Belgium d Department of Environmental Science, Aarhus University, Frederiksborgvej 399, PO Box 358, DK-4000 Roskilde, Denmark b
art ic l e i nf o
a b s t r a c t
Article history: Received 1 April 2016 Received in revised form 8 July 2016 Accepted 11 July 2016
To better elucidate the potential immune-related health effects of exposure to environmentally persistent organic pollutants (POP), such as polychlorinated biphenyls (PCBs) and perfluoroalkyl substances (PFASs), in ringed seals (Pusa hispida), a sentinel Arctic species, we assessed 1) associations between mitogen-induced lymphocyte proliferation and in vivo tissue contaminant burdens, and 2) the concentration-response effects of in vitro exposure to PFASs and PCB congeners on mitogen-induced lymphocyte proliferation. Upon in vitro contaminant exposure, the non-coplanar PCB congeners CB 138, 153, and 180, but not the coplanar CB 169, significantly reduced lymphocyte proliferation between 10 and 20 mg g 1 ww. The respective in vitro EC50 values for these congeners were 13.3, 20.7, 20.8, and 54.6 mg g 1 ww. No modulation of lymphocyte proliferation was observed upon in vitro exposure to two individual PFASs, perfluorooctane sulphonic acid (PFOS) and perfluorooctanoic acid (PFOA), at concentrations up to 1000 ng g-1. In addition, no significant correlations were found between lymphocyte proliferation and any blood or blubber contaminant measured. Taken together, these data suggest this population of ringed seals is not currently at high risk of altered lymphocyte proliferation from exposure to the POPs or PFASs in this study. & 2016 Published by Elsevier Inc.
Keywords: Ringed seal PCB PFAS Immunotoxicology Lymphocyte proliferation
1. Introduction Environmentally persistent organic pollutants (POP), such as polychlorinated biphenyls (PCBs) and perfluoroalkyl substances (PFASs), have been detected in the tissues of numerous marine mammals inhabiting the Arctic, an area remote from major sources of production and release (Letcher et al., 2010; Riget et al., 2013; Vorkamp et al., 2011). Effects of these chemicals on the immune system have been well characterized in controlled laboratory experiments, and have been suggested in several wildlife species as well (Letcher et al., 2010). The immune system is a complex network of cells with diverse functions, communicating through a wide array of messenger molecules, which has evolved to protect the host from potentially pathogenic agents including viruses, bacteria, and parasites (Kuby, 2012). An immunomodulatory effect on any number of essential functions within this complex system is likely to have adverse consequences n
Corresponding author. E-mail address: [email protected] (M. Levin).
http://dx.doi.org/10.1016/j.envres.2016.07.013 0013-9351/& 2016 Published by Elsevier Inc.
on an organism's survival and fitness, and its potential to handle several environmental stressors. Indeed, contaminant-induced immunomodulation has been suggested as a contributing factor in several morbillivirus-associated marine mammal mass mortality events (Di Guardo et al., 2005; Ross et al., 1996; Van Loveren et al., 2000). The persistence of PCBs in the environment is due in part to long half-lives and high lipophilicity, leading to bioaccumulation and biomagnification in food webs. The lack of ortho-positioned chlorine atoms allows a congener to produce a coplanar (non-ortho) configuration similar to that of the most toxic of the halogenated aromatic hydrocarbons, 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Congeners with two or more chlorines within the ortho positions (di-ortho PCBs) experience steric hindrance resulting in their aromatic rings being offset into separate planes (non-coplanar PCBs). The role of the aryl hydrocarbon receptor (AhR) in dioxin and dioxin-like (co-planar) PCBs immunotoxicity has been well established in laboratory rodent studies, and is linked in particular to effects on developing lymphoid organs and subsequent cellular and humoral defects (Lawrence and Nancy, 2006).
M. Levin et al. / Environmental Research 151 (2016) 244–250
However, non-coplanar PCBs, with little to no affinity for the AhR, were shown to modulate harbor seal (Phoca vitulina), dolphin (Tursiops truncatus), and beluga (Delphinapterus leucas) immune functions upon either in vitro or in vivo exposure (Levin et al., 2005a, 2004, 2005b; Mori et al., 2008, 2006) and the toxicity pathways are therefore likely very different from those induced by the AhR (Levin et al., 2007b). PFASs, such as perfluorooctanoic acid (PFOA) and perfluorooctane sulphonic acid (PFOS), have physical and chemical properties that make them ideal surfactants (Lau et al., 2004). Although the manufacturing process results mostly in the production of perfluorooctane sulfonyl fluoride (POSF), PFOS is a common end-stage metabolite and product (Lau et al., 2004). PFOS and PFOA are readily absorbed through the gastro-intestinal tract, distributed in serum and liver, and poorly excreted or metabolized (Seacat et al., 2002). The immunotoxic effect of PFASs was recently reviewed (Dewitt et al., 2012) and involves activation of the alpha isotype of peroxisome proliferator-activated receptors (PPARs) in laboratory rodents, a ligand-activated transcription factor that regulates gene expression (DeWitt et al., 2009). However, some immunotoxic effects may occur independently of the receptor (DeWitt et al., 2009). PFAS are transported to the Arctic via pathways including airborne precursors (e.g. fluorinated sulphonamides) followed by atmospheric oxidative transformations, as well as by direct long-range oceanic transport (Schenker et al., 2008). Unlike the lipophilic legacy PCBs, PFASs bind to blood proteins and bioaccumulate mainly in liver, kidneys and bile secretions (Jones et al., 2003) and have been measured in Greenland ringed seals, as well as polar bears (Riget et al., 2013). To better elucidate the immune effects of POPs in an Arctic sentinel species, we collected live lymphocytes from ringed seals during Inuit subsistence hunting in Scoresby Sound, East Greenland. The ringed seal serves as a key species in the Arctic Monitoring and Assessment Programme (AMAP) already for several decades due to its widespread distribution, its year-round presence and its ecological and nutritional importance within the entire Arctic area (AMAP, 2004). We aimed to test whether select POPs are immunotoxic to free-ranging ringed seals from East Greenland upon either in vitro or in vivo exposure. More specifically, we aimed at assessing 1) associations between mitogen-induced lymphocyte proliferation and in vivo tissue contaminant burdens, and 2) the concentration-response effects of in vitro exposure to PFASs and PCB congeners on mitogen-induced lymphocyte proliferation. The mitogen-induced lymphocyte proliferation assay measures the ability of lymphocytes to proliferate in response to a polyclonal stimulation. For marine mammal immunotoxicity studies, lymphocyte proliferation is by far the most commonly used immune assay, with studies using different organic pollutants and metals, predominantly reporting immunosuppressive effects despite the many differences in study design and animal life history (Desforges et al., 2016).
2. Materials and methods 2.1. Sample collection Nine ringed seals were collected between 22 February and 5 March 2012 during the legal local subsistence hunt in the area around Ittoqqorrtoormiit, East Greenland (70° 24′ 0′′ N, 21° 58′ 0′′ W). Seals were either shot while hauled out on ice or in the water or caught in nets under the sea ice. Sections of mesenteric lymph nodes were collected from harvested ringed seals within 10– 60 min of death, under Aarhus University's permits and supervision. Sections of lymph nodes were cut into 5 5 mm pieces using a sterile scalpel, and were placed in cold Dulbecco’s modified
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Eagle’s medium (DMEM) culture medium prior to preparing single cell suspensions for cryopreservation in the vapour phase of liquid nitrogen, as previously described for field samples of marine mammals (De Guise et al., 1996). Full depth blubber samples, sections of most major organs, and blood were also collected for contaminant analysis and/or other research programs. Tissues were imported into Denmark and further distributed to send to various collaborators including the University of Connecticut (United States; National Marine Fisheries Service Permit #10081637-03) and the University of Antwerp (Belgium). All tissue samples were stored at 20 °C while lymph node cells were kept in liquid nitrogen until analytical processing. 2.2. Analysis of blubber POP concentrations The analytical protocol was based on previous methods (Weijs et al., 2009) and allowed for the analysis for 42 PCB congeners (CB 18, 28, 31, 44, 47, 49, 52, 66, 70, 74, 87, 95, 99, 101, 105, 110, 118, 128, 132, 138, 146, 149, 151, 153, 156, 158, 170, 171, 172, 174, 177, 180, 183, 187, 194, 195, 196, 199, 203, 205, 206 and 209), several organochlorine pesticides (OCPs), amongst which dichlorodiphenyltrichloroethane (p,p′- and o,p′-DDT) and its metabolites (p,p′- and o,p ′-DDD and p,p′- and o,p′-DDE), hexachlorobenzene (HCB), cis- and trans-chlordanes (CN and TN) and its metabolites oxychlordane (OxC) and cis- and trans-nonachlor (CN and TN), and several PBDE congeners (BDE 28, 47, 49, 99, 100, 153, 154, and 183). An accurately weighed blubber subsample (median: 0.14 g; range: 0.09– 0.23 g) was homogenised and mixed with anhydrous sodium sulphate, spiked with internal standards (CB 143 and BDE 77), and extracted for 2 h with hexane: acetone (3:1, v-v) in an automated hot Soxhlet extractor (Büchi Flawil, CH). The lipid content was determined gravimetrically on 1/10th of the extract (after 1 h at 100 °C) prior to clean up on 8 g acidified silica (44% sulphuric acid). PBDEs, CHLs, and higher PCBs were measured with an Agilent 6890–5973 gas chromatograph coupled to a mass spectrometer system (GC–MS), equipped with a 30 m x 0.25 mm x 0.25 mm DB-5 capillary column (J&W Scientific, Folsom, CA, USA) and operated in electron capture negative ionisation mode. Lower PCBs, DDXs (DDT, DDD, DDE), and HCB were quantified using an identical GC–MS operated in electron ionisation mode and equipped with a 25 m x 0.22 mm x 0.25 mm HT-8 capillary column (SGE, Zulte, Belgium). Procedural blanks, analyzed simultaneously with every batch of seven samples, were consistent (relative standard deviation o 30%) and therefore the mean value was calculated for each compound and subtracted from the sample values. The limit of quantification (MLOQ) was set at 3*standard deviation (SD) of the mean blank measurements, or, for analytes not detected in procedural blanks, set at signal to noise ratio (S: N) ¼10:1. Mean 7SD recoveries of the internal standards CB 143 and BDE 77 were 86 76% and 93 710%, respectively. The analytical procedure was validated through the analysis of certified material SRM 1945 for which deviations from certified values were less than 10%. This quality control scheme has been positively assessed through participation in the NIST/NOAA Intercomparison Programme for Organic and Inorganic Contaminants in Marine Mammal Tissues. 2.3. Analysis of serum PFAS concentrations The extraction method for PFAS in serum was based on solid phase extraction (SPE) as previously described (Keller et al., 2010). Before extraction, the samples (100 ml volume) were spiked with 1 ng of each of the following 13C-labelled internal standards: 13C2PFHxA, 13C4-PFOA, 13C5-PFNA, 13C2-PFDA, 13C2-PFUnA, 13C2-PFDoA, 13 C8-PFOSA, 13C4-PFHxS, 13C4-PFOS, all obtained from Wellington Laboratories (Guelph, ON, Canada). Instrumental analysis was
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performed by liquid chromatography-tandem mass spectrometry (LC-MS-MS) with electrospray ionisation (ESI) in negative mode. The samples were extracted and analyzed in batches together with a procedural blank and two control samples, represented by an aliquot of test material previously analyzed in a ring test and for which assigned values for PFOS and PFOA concentrations are available. The detection limits ranged from 0.02 to 0.7 ng mL 1. As an additional quality control procedure, the performance of analytical method is regularly tested three times a year by participating in the Arctic Monitoring and Assessment Programme (AMAP) Ring Test for Organic Pollutants in Human Serum organized by the Institute Nationale de Santé Publique du Québec. Detection limit and precision of the analytical method were determined using spiked rabbit serum (Sigma Aldrich). 2.4. Preparation of immune cells Cryovials containing lymph node cell suspensions were quickly thawed in a 37 °C water bath and the contents were transferred to a conical tube containing 10 times the volume of warm DMEM supplemented with 1 mM sodium pyruvate, 100 mM non-essential amino acids, 25 mM HEPES, 2 mM L-glutamine, 100 U mL 1 penicillin, 100 mg mL 1 streptomycin and 0.25 mg mL 1 Fungizone (all obtained from Life, Grand Island, NY, USA), along with 10% fetal bovine serum (Hyclone, Logan, UT, USA), thereafter referred to as complete DMEM. Tubes were centrifuged at 220 g for 10 min at room temperature, and the cell pellet was re-suspended in complete DMEM. After thawing and washing once in DMEM, the cell suspension was passed through a sterile pipette containing sterile nylon wool to remove any residual connective tissue. To remove dead cells, the cell suspension was processed using the MACS Dead Cell Removal Kit (Miltenyi Biotec, Auburn, CA, USA), as per manufacturer's instructions. After collecting and washing the live cells, as confirmed by trypan blue staining (viability above 95%), the cell suspension was adjusted to 2 106 cells mL 1. 2.5. Mitogen-induced lymphocyte proliferation Mitogen-induced lymphocyte proliferation, without in vitro chemical exposure, was evaluated by methods routinely performed in our laboratory (Levin et al., 2007a; Mori et al., 2006). Lymphocytes were plated in 96 well flat bottom tissue culture plates, in triplicate, and incubated at 37 °C with 5% CO2 for a total of 66 h with the T cell mitogen concanavalin A (ConA), the B cell mitogen lipopolysaccharide (LPS), or complete DMEM medium alone (no mitogen, unstimulated). Con A and LPS were used at the sub-optimal concentrations of 0.1 mg mL 1 and 0.05 mg mL 1, respectively. Sub-optimal concentrations were used as they proved more sensitive in detecting immunotoxicity (Mori et al., 2006). Lymphocyte proliferation was reported as the stimulation index (SI ¼ mitogen/no mitogen). Mitogen-induced lymphocyte proliferation with in vitro chemical exposure was evaluated as above, though with addition of chemical during incubation. For the PCB congeners, the concentrations chosen (range 0.5–20 mg g 1 ww in culture media) were within effect levels seen in previous in vitro exposure studies (De Guise et al., 1998; Mori et al., 2008; Mori et al., 2006) and at concentrations 2 to 10 fold higher in plasma of ringed seals from the Baltic and Svalbard areas (maximum values of 0.185 and 0.260 mg g 1 ww plasma, respectively) (Routti et al., 2008). When converted to a lipid weight, the ΣPCB concentration range in the Baltic and Svalbard seals were comparable to the ΣPCB concentration range found in the blubber of seals in the current study. For PFOA and PFOS, the concentrations chosen (range 1– 1000 ng g 1 ww) were also within and above the concentrations measured in ringed seal serum in this study (range 0.6–3.1 ng g 1
ww PFOA; range 21.1–196.6 ng g1 ww PFOS). The following chemicals and final concentrations were tested: PCB congeners 138, 153, 180 and 169 (ULTRA Scientific, N. Kingstown, RI, USA) were tested at 0 (unexposed control), 0.5, 1.0, 2.5, 5, 10, 15, and 20 mg g 1. PFOA (Sigma-Aldrich, St. Louis, MO, USA) was tested at 0 (unexposed control), 1, 3, 10, 30, 100, 300, 1000 ng g 1. PFOS (Sigma-Aldrich, St. Louis, MO, USA) was tested at 0 (unexposed control), 1, 3, 10, 30, 100, 300 ng g 1. Both PFASs were tested using the mitogens ConA and LPS, while the PCB congeners were tested with ConA only due to limited lymphocytes available for testing. 2.6. Statistical analysis Since immune cells from the same individual were divided into several chemical exposure groups (in vitro exposure), one-way repeated measures ANOVA with Dunnett's test, using the SigmaStat 3.5 (Jandel Scientific, San Rafael, CA) software, was used to compare the effects at the different in vitro concentrations to unexposed control cells. This was done to determine what concentrations significantly affected proliferation, and determine the lowest concentration that was significantly different from control (LOEC: Lowest Observed Effective Concentration). Concentrationresponse curves were generated by plotting the effects of the different treatment groups as a percentage of the unexposed control cells, and determining mathematically what concentration resulted in a 50% change (decrease or increase) in response (EC50). Correlation analyses using Pearson product-moment correlation were performed using SPSS 16.0 (IBM, New York, USA). For all tests, po 0.05 was used for statistical significance. A t-test was used to compare contaminant concentrations between males/females and adults/subadults.
3. Results Table 1 summarizes the biological and contaminant data for the major POPs (ΣPCBs, ΣDDTs, ΣHCHs, ΣPBDEs, HCB) and PFASs (PFOS and PFOA) for each ringed seal. Full details of all blubber/ serum concentrations are provided in the supplementary information (Table S1). Nine adult or sub-adult seals were captured in total, six males and three females, with all but one (ID 43166) in good body condition. Contaminant concentrations were neither significantly different between males and females nor between adults and sub-adults (p4 0.05 for all contaminants). No significant correlations existed between lymphocyte proliferation (with and without mitogen) and any tissue contaminant concentration in the nine ringed seals (p 40.05 for all comparisons; data not shown). For the in vitro exposure tests, exposure to the non-coplanar CB 138, 180 and 153 significantly reduced ConA stimulated lymphocyte proliferation at 10, 15 and 20 mg g 1 respectively (Fig. 1 and Fig. S1). Furthermore, the relative magnitude of reduction in lymphocyte proliferation was greater for CB 138 than CB 153 and 180. In contrast, the coplanar CB 169 did not significantly modulate ConA stimulated lymphocyte proliferation at any concentration tested (p 40.05; Fig. 1 and Fig. S1). EC50 values were determined at 13.3, 20.7, 20.8, and 54.6 mg g 1 for CB 138, 153, 180, and 169, respectively (Fig. 1). EC50 values 420 were extrapolated beyond the 0–20 range based on the best fit curve. Unstimulated lymphocyte proliferation was not significantly modulated by any PCB congener or concentration tested (p 40.05; data not shown). ConA- and LPS-induced lymphocyte proliferation was not significantly modulated by any concentration of PFOA or PFOS tested (p 40.05; Fig. S2).
M. Levin et al. / Environmental Research 151 (2016) 244–250
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Table 1 Biological and contaminant data for nine East Greenland ringed seals. Serum (ng mL 1)
Blubber (ng g 1 lw)
Seal ID Sex Age
Zoological length (cm)
Blubber thickness (mm)
Weight category
lipid (%) PFOS
PFOA
ΣPCBs
ΣDDXs ΣHCHs ΣPBDEs HCB
43181 43189 43187 43184 43183 43188 43166 43192 43190
122.5 122 110 99 121 104 107 115 110
72 60 40 48 60 45 30 57 40
Very fat Fat Medium Medium Fat Medium Thin Fat Medium
85.4 84.6 88.8 78.5 90.7 81.8 82.7 97.8 92.5
25.6 21.1 31.6 196.6 51.6 49.3 48.2 45.7 46.0
1.1 1.3 0.6 3.1 1.9 1.2 2.6 2.6 1.6
349 349 349.4 349.6 483.5 545 569.2 663 2852.5
264.6 342.2 322.4 273.2 382.9 426.6 521.6 568.9 2761.8
15.7 22.6 28.5 38.6 32.7 37.8 25.1 28.5 51.7
9.1 10.8 10.6 9.6 13.4 14.6 18.8 18.9 49.4
9.2 10.4 10.3 8.9 11.1 6.7 8.6 6.4 12.3
Mean SD Min Max
87 6 78.5 97.8
57.3 53.4 21.1 196.6
1.8 0.8 0.6 3.1
723.4 807.0 349 2852.5
651.6 798.2 264.6 2761.8
31.2 10.5 15.7 51.7
17.2 12.6 9.1 49.4
9.3 1.9 6.4 12.3
F M M F M M M F M
Adult Adult Adult Subadult Adult Subadult Subadult Adult Subadult
125
*
***
***
10
15
20
% control
100
75
50
25
0 0
5
concentration (µg g-1)
Fig. 1. Concentration-response curves for ConA-induced lymphocyte proliferation upon in vitro exposure to increasing concentrations of CB 138, 153, 180, and 169. Data for each congener are expressed as a percent of the unexposed control. EC50 ¼ effective concentration that reduced lymphocyte proliferation by 50%. *, p o0.05 compared to unexposed control for each individual congener (one-way RM ANOVA). Error bars are þ/ 1 SD.
4. Discussion This study utilized blubber, blood, and lymph node immune cells from freshly harvested ringed seal to evaluate the effects of two major classes of environmentally persistent contaminants, i.e. PCBs and PFASs, upon either in vitro or natural exposure, on the immune system. No significant associations were found between mitogen-induced lymphocyte proliferation and tissue contaminant concentrations. However, in vitro exposure to the noncoplanar PCB congeners 138, 153, and 180 significantly reduced lymphocyte proliferation at 10 mg g 1 or above. The lack of significant relationships between lymphocyte proliferation and tissue contaminant concentrations was not surprising, as the concentrations of tissue contaminants were below concentrations shown to modulate immune functions in other marine mammals and laboratory animals. Serum PFOS (range 21–197 ng mL 1 ww) measured in this study was within the range of concentrations previously reported between 1982 and 2010 in East (liver, 20–352 ng mL 1 ww) and West (liver, 12 and 397 ng mL 1 ww) Greenland ringed seals (Riget et al., 2013). Serum PFOA (range 0.6–3 ng mL 1 ww) measured in this study was within the range of concentrations previously reported between 1982 and 2010 in East (liver, below minimum
detection levels to 2.2 ng mL 1 ww) and West (liver, below minimum detection levels to 2.4 ng mL 1 ww) (Riget et al., 2013) Greenland ringed seals. Serum PFOS (mean 57.3 ng mL 1) and serum PFOA (mean 1.8 ng mL 1) measured in the current study were below the lowest observed adverse effect level (LOAEL) in serum that induced immunotoxicity (notably T-cell-dependent IgM antibody response) in laboratory mice (Dewitt et al., 2012), 92 ng mL 1 for PFOS and 74,000 ng mL 1 for PFOA, with one exception, seal ID 43184, which had a serum PFOS concentration of 196.6 ng mL 1. Blubber ΣPCBs (range 349–2852 ng g 1 lw) measured in this study was within the range of concentrations previously reported in East Greenland ringed seals in 2004 (ΣPCB10 mean 605 þ/ 240 ng g 1 lw, ) (Riget et al., 2006) and above the concentration range previously reported for West Greenland ringed seals between 1994 and 2004 (ΣPCB10 133–332 ng g 1) (Vorkamp et al., 2008). TheΣPCB concentrations in this study were below the threshold for physiological effects (1370 ng g 1 lw blubber) as previously reported for aquatic mammals, including ringed seals (Brown et al., 2014; Kannan et al., 2000). Concentrations were, however, within range of the effect threshold for lymphocyte proliferation recently established for marine mammals, as well as for seals specifically. In their meta-analysis of all available marine mammal immunotoxicology studies (including in vitro and in vivo data), Desforges et al. (Desforges et al., 2016) found a sigmoidal dose-dependent decrease in lymphocyte proliferation, with seal threshold effect levels of ΣPCBs of 8 70.01 ng g 1, respectively. Concentrations in this study were also within range of estimated critical body residues (CBR) for immunosuppression in polar bears (Ursus maritimus), 469 ng g 1 lw blubber (Dietz et al., 2015). CBR is defined as the lowest observed total body concentration of a contaminant, which is associated with the occurrence of adverse toxic effects. Although marine mammal CBRs have only been calculated for polar bears, they may serve as a reference for other marine mammals until species-specific CBRs are defined. ΣDDX (range 254–2762 ng g 1) measured in this study was found within a similar concentration range previously reported in East Greenland ringed seals (adult, 610–2460 ng g 1 lw; subadult, 360–1070 ng g 1 lw) (Vorkamp et al., 2004) and the concentration range previously reported in West Greenland ringed seals between 1999 and 2004 (190–549 ng g 1 lw) (Vorkamp et al., 2008). Blubber p,p′-DDT measured in this study was below concentrations found in blubber of harbor seals fed either herring from the ‘uncontaminated’ Atlantic ocean or ‘contaminated’ Baltic Sea (mean 306 and 2448 ng g 1 lw, respectively). Only seals fed the
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Baltic Sea herring had a significant reduction in mitogen-induced T lymphocyte proliferation (de Swart et al., 1996), but this was related to TEQ levels in blubber, and not only DDT specifically. In addition, ΣDDXs in this study (mean 652 ng g 1 lw blubber) was also 10-fold below the estimated CBR for immunosuppression in polar bears, 6502 ng g 1 lw blubber (Dietz et al., 2015). Blubber ΣHCHs (range 16–52 ng g 1 lw) measured in this study was similar to the concentration range previously reported in East Greenland ringed seals (adult, 53–92 ng g 1 lw; subadult, 29–120 ng g 1 lw) (Vorkamp et al., 2004) and similar to the median concentration previously reported in West Greenland ringed seals in 2006 (47 ng g 1 lw) (Vorkamp et al., 2008). Blubber ΣHCHs (mean 31.2 ng g 1 lw) measured in this study was nearly 42-fold below the estimated CBR for immunosuppression in polar bears, 1302 ng g 1 lw blubber (Dietz et al., 2015). Blubber ΣPBDEs (range 9–49 ng g 1 lw) measured in this study was similar to the mean concentration previously reported in East Greenland ringed seals in 2004 (mean 32.5 þ/ 13.8 ng g 1 lw) (Riget et al., 2006) and slightly above the median range previously reported in West Greenland ringed seals (2–8 ng g 1 lw) (Vorkamp et al., 2008). Blubber ΣPBDEs (mean 17.2 ng g 1 lw) measured in this study were below concentrations ( 18 106 ng g 1 lw liver) found to modulate immune functions (i.e. antibody production) in experimentally exposed mink (Martin et al., 2007). Blubber HCB (range 6–12 ng g 1 lw) measured in this study was similar to the concentration range previously reported in East Greenland ringed seals (adult, 9–20 ng g 1 lw; subadult, 8– 27 ng g 1 lw). To date, relationships between HCB and changes in marine mammal immune functions have not been documented (Dupont et al., 2013). Overall, most tissue contaminant concentrations were below concentrations found to induce immunotoxicity in other marine mammals and laboratory animals. The only exception are PCBs where blubber levels appear to fall within range of effect thresholds calculated for marine mammal lymphocyte proliferation (Desforges et al., 2016); thus concentrations in this study may be bordering or just above the no observable effect level (NOEL). A larger sample size is however necessary to observe an in vivo correlation between immune functions and PCB concentrations. To further examine the effects of contaminants on the immune system of ringed seals, in vitro exposure experiments were conducted with four individual PCB congeners, CB 138, 153, 180 and 169, and two PFSAs, PFOA and PFOS. We chose in vitro concentrations, on a weight wet basis (mg g 1 or ng g 1), that were within and above those found upon natural in vivo exposure (in serum or plasma), also on a weight wet basis, in other populations of ringed seals for which published data exist (see Methods). This allowed us to use an environmental relevant concentration range to assess the immunotoxic potential of these chemicals on circulating immune cells using an in vitro approach. In vitro experiments allow for the detection of the direct effects of contaminants on immune functions. When performing the in vitro experiments, cells from the same individual serve as the unexposed control for that individual; therefore, it takes into account any previous POP exposure and other confounding factors such as age and sex. Therefore, any modulation of immune function is the result of the actual in vitro exposure. In addition, Desforges et al. (2016) found that for lymphocyte proliferation, dose response curves for in vitro and in vivo lymphocyte proliferation vs. pollutant concentrations overlapped very well, suggesting that the in vitro approach can be suggestive of the outcome of in vivo exposure to similar concentrations. Importantly, the use of in vitro concentration–response experiments allows the testing of a wider range of concentrations, especially at the higher end of the range, which animals may experience if contaminant burdens continue to
increase. This may significantly improve wildlife risk assessment. In addition, the use of in vitro testing allows one to assess and compare the immunotoxic potential of individual contaminants at the cell level and compare among different species (Levin et al., 2007b, 2007c, 2005b; Mori et al., 2006). In this study, there was a concentration dependent suppression of T lymphocyte proliferation at concentrations of 10 mg g 1 or higher for non-coplanar PCBs. Similar immunomodulatory effects were documented for non-coplanar PCBs in B6C3F1 mouse, managed-care beluga whales, and human T lymphocyte proliferation under comparable in vitro conditions (Leibrecht, 2007). For example, suppressed lymphocyte function was found in mice for PCB congeners 138, 153 and 180 and in beluga for CB 138, however, lymphocyte proliferation increased in beluga exposed to CB 153 and 169 and in humans for all four congeners. Not only is the directionality of in vitro lymphocyte response variable between species, but the effect concentrations and thresholds differ as well. For example, human lymphocytes were most sensitive to CB 138 with LOEC and EC50 values of 5 and 6.9 mg g 1, compared to 10.0 and 13.3 mg g 1 for ringed seals, 15.0 and 19.4 mg g 1 for mouse, and 18.7 and 420 mg g 1 for beluga. Similar to natural exposure, there was no modulation of mitogen-induced lymphocyte proliferation upon in vitro exposure to PFOA or PFOS. For PFOS, the estimated CBR for immunotoxic effects in polar bears was 9356 ng g 1 ww (Dietz et al., 2015), with no value available for PFOA. The LOAEL for PFOS that induced immunotoxicity in laboratory mice, 92 ng mL 1 serum, was within the range of in vitro concentrations tested here, while the LOAEL for PFOA, 74,000 ng mL 1 serum, was above the range tested here (Dewitt et al., 2012).
5. Conclusions Taken together, these data suggest this population of ringed seals is not currently at high risk for immunotoxic effects upon POP or PFAS exposure examined in this study. However, under stressful conditions such as nutritional deficiency and reproduction, marine mammals may undergo periods of blubber metabolism and fat mobilization such that blubber PCBs would concurrently mobilize and reach higher levels in circulation (Debier et al., 2006; Yordy et al., 2010). Thus, the higher concentrations in blood may reach levels demonstrated to temporarily modulate immune functions, potentially increasing an individual's risk for immunomodulation during already stressful periods. In addition, these data provide further ‘weight of evidence’ that PCBs, especially the non-coplanar congeners (CB 138, 153, and 180), are potentially immunotoxic in marine mammals, although difference among species susceptibility must be considered as previously documented (Mori et al., 2006). At current levels, PFASs appear not to be of concern on some aspects of the immune system. Importantly, measuring other aspects of immune functions, such as key innate and adaptive immune functions, including phagocytosis, respiratory burst, and natural killer (NK) cell activity, will provide a broader and clearer picture on the effects of contaminants on the overall health of ringed seals, as well as other top Arctic predators. Ideally, similar studies as the present should be conducted for marine mammals at risk including seals from the Baltic seals and other marine mammals foraging at higher trophic levels, including e.g. harp seals (Pagophilus groenlandicus), hooded seals (Cystophora cristata), toothed whales like narwhals (Monodon monoceros), beluga whales, pilot whales (Globicephala melas) and killer whales (Orcinus orca) as well as polar bears (Ursus maritimus).
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Funding sources Major funding for this project was provided by the DANCEA (Danish Cooperation for Environment in the Arctic) and ARC (Arctic Research Centre). Additional travel support (for Levin) was provided by the University of Connecticut Research Foundation and the University of Connecticut's Interdisciplinary Research and Training Initiative on Coastal Ecosystems and Human Heath (IRICH). Funding travel support was also provided by the Norwegian University of Science and Technology, N-7491, Trondheim, Norway.
Acknowledgements We thank local subsistence Scoresby Sound hunters for helping with the sample collection at Kap Tobin in February–March 2012, as well as colleagues from four countries (Denmark, Norway, Canada, United States) as part of the AURORAE project. Major funding for this project was provided by the DANCEA (Danish Cooperation for Environment in the Arctic) and ARC (Arctic Research Centre). Additional travel support (for Levin) was provided by the University of Connecticut Research Foundation and the University of Connecticut's Interdisciplinary Research and Training Initiative on Coastal Ecosystems and Human Heath (I-RICH). Funding travel support was also provided by the Norwegian University of Science and Technology, N-7491, Trondheim, Norway.
Appendix A. Supplementary material Supplementary data associated with this article can be found in the online version at http://dx.doi.org/10.1016/j.envres.2016.07. 013.
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Immunomodulatory effects of exposure to polychlorinated biphenyls and perfluoroalkyl acids in East Greenland ringed seals (Pusa hispida) Milton Levin a,n, Erika Gebhard a, Lindsay Jasperse a, Jean-Pierre Desforges b, Rune Dietz b, Christian Sonne b, Igor Eulaers b, Adrian Covaci c, Rossana Bossi d, Sylvain De Guise a a
Department of Pathobiology and Veterinary Science, University of Connecticut, 61 North Eagleville Road, Storrs, CT 06269-3089, United States Department of Bioscience, Arctic Research Centre (ARC), Aarhus University, Frederiksborgvej 399, PO Box 358, DK-4000 Roskilde, Denmark c Toxicological Centre, Department of Pharmaceutical Sciences, University of Antwerp, Universiteitsplein 1, BE-2610 Wilrijk, Belgium d Department of Environmental Science, Aarhus University, Frederiksborgvej 399, PO Box 358, DK-4000 Roskilde, Denmark b
art ic l e i nf o
a b s t r a c t
Article history: Received 1 April 2016 Received in revised form 8 July 2016 Accepted 11 July 2016
To better elucidate the potential immune-related health effects of exposure to environmentally persistent organic pollutants (POP), such as polychlorinated biphenyls (PCBs) and perfluoroalkyl substances (PFASs), in ringed seals (Pusa hispida), a sentinel Arctic species, we assessed 1) associations between mitogen-induced lymphocyte proliferation and in vivo tissue contaminant burdens, and 2) the concentration-response effects of in vitro exposure to PFASs and PCB congeners on mitogen-induced lymphocyte proliferation. Upon in vitro contaminant exposure, the non-coplanar PCB congeners CB 138, 153, and 180, but not the coplanar CB 169, significantly reduced lymphocyte proliferation between 10 and 20 mg g 1 ww. The respective in vitro EC50 values for these congeners were 13.3, 20.7, 20.8, and 54.6 mg g 1 ww. No modulation of lymphocyte proliferation was observed upon in vitro exposure to two individual PFASs, perfluorooctane sulphonic acid (PFOS) and perfluorooctanoic acid (PFOA), at concentrations up to 1000 ng g-1. In addition, no significant correlations were found between lymphocyte proliferation and any blood or blubber contaminant measured. Taken together, these data suggest this population of ringed seals is not currently at high risk of altered lymphocyte proliferation from exposure to the POPs or PFASs in this study. & 2016 Published by Elsevier Inc.
Keywords: Ringed seal PCB PFAS Immunotoxicology Lymphocyte proliferation
1. Introduction Environmentally persistent organic pollutants (POP), such as polychlorinated biphenyls (PCBs) and perfluoroalkyl substances (PFASs), have been detected in the tissues of numerous marine mammals inhabiting the Arctic, an area remote from major sources of production and release (Letcher et al., 2010; Riget et al., 2013; Vorkamp et al., 2011). Effects of these chemicals on the immune system have been well characterized in controlled laboratory experiments, and have been suggested in several wildlife species as well (Letcher et al., 2010). The immune system is a complex network of cells with diverse functions, communicating through a wide array of messenger molecules, which has evolved to protect the host from potentially pathogenic agents including viruses, bacteria, and parasites (Kuby, 2012). An immunomodulatory effect on any number of essential functions within this complex system is likely to have adverse consequences n
Corresponding author. E-mail address: [email protected] (M. Levin).
http://dx.doi.org/10.1016/j.envres.2016.07.013 0013-9351/& 2016 Published by Elsevier Inc.
on an organism's survival and fitness, and its potential to handle several environmental stressors. Indeed, contaminant-induced immunomodulation has been suggested as a contributing factor in several morbillivirus-associated marine mammal mass mortality events (Di Guardo et al., 2005; Ross et al., 1996; Van Loveren et al., 2000). The persistence of PCBs in the environment is due in part to long half-lives and high lipophilicity, leading to bioaccumulation and biomagnification in food webs. The lack of ortho-positioned chlorine atoms allows a congener to produce a coplanar (non-ortho) configuration similar to that of the most toxic of the halogenated aromatic hydrocarbons, 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Congeners with two or more chlorines within the ortho positions (di-ortho PCBs) experience steric hindrance resulting in their aromatic rings being offset into separate planes (non-coplanar PCBs). The role of the aryl hydrocarbon receptor (AhR) in dioxin and dioxin-like (co-planar) PCBs immunotoxicity has been well established in laboratory rodent studies, and is linked in particular to effects on developing lymphoid organs and subsequent cellular and humoral defects (Lawrence and Nancy, 2006).
M. Levin et al. / Environmental Research 151 (2016) 244–250
However, non-coplanar PCBs, with little to no affinity for the AhR, were shown to modulate harbor seal (Phoca vitulina), dolphin (Tursiops truncatus), and beluga (Delphinapterus leucas) immune functions upon either in vitro or in vivo exposure (Levin et al., 2005a, 2004, 2005b; Mori et al., 2008, 2006) and the toxicity pathways are therefore likely very different from those induced by the AhR (Levin et al., 2007b). PFASs, such as perfluorooctanoic acid (PFOA) and perfluorooctane sulphonic acid (PFOS), have physical and chemical properties that make them ideal surfactants (Lau et al., 2004). Although the manufacturing process results mostly in the production of perfluorooctane sulfonyl fluoride (POSF), PFOS is a common end-stage metabolite and product (Lau et al., 2004). PFOS and PFOA are readily absorbed through the gastro-intestinal tract, distributed in serum and liver, and poorly excreted or metabolized (Seacat et al., 2002). The immunotoxic effect of PFASs was recently reviewed (Dewitt et al., 2012) and involves activation of the alpha isotype of peroxisome proliferator-activated receptors (PPARs) in laboratory rodents, a ligand-activated transcription factor that regulates gene expression (DeWitt et al., 2009). However, some immunotoxic effects may occur independently of the receptor (DeWitt et al., 2009). PFAS are transported to the Arctic via pathways including airborne precursors (e.g. fluorinated sulphonamides) followed by atmospheric oxidative transformations, as well as by direct long-range oceanic transport (Schenker et al., 2008). Unlike the lipophilic legacy PCBs, PFASs bind to blood proteins and bioaccumulate mainly in liver, kidneys and bile secretions (Jones et al., 2003) and have been measured in Greenland ringed seals, as well as polar bears (Riget et al., 2013). To better elucidate the immune effects of POPs in an Arctic sentinel species, we collected live lymphocytes from ringed seals during Inuit subsistence hunting in Scoresby Sound, East Greenland. The ringed seal serves as a key species in the Arctic Monitoring and Assessment Programme (AMAP) already for several decades due to its widespread distribution, its year-round presence and its ecological and nutritional importance within the entire Arctic area (AMAP, 2004). We aimed to test whether select POPs are immunotoxic to free-ranging ringed seals from East Greenland upon either in vitro or in vivo exposure. More specifically, we aimed at assessing 1) associations between mitogen-induced lymphocyte proliferation and in vivo tissue contaminant burdens, and 2) the concentration-response effects of in vitro exposure to PFASs and PCB congeners on mitogen-induced lymphocyte proliferation. The mitogen-induced lymphocyte proliferation assay measures the ability of lymphocytes to proliferate in response to a polyclonal stimulation. For marine mammal immunotoxicity studies, lymphocyte proliferation is by far the most commonly used immune assay, with studies using different organic pollutants and metals, predominantly reporting immunosuppressive effects despite the many differences in study design and animal life history (Desforges et al., 2016).
2. Materials and methods 2.1. Sample collection Nine ringed seals were collected between 22 February and 5 March 2012 during the legal local subsistence hunt in the area around Ittoqqorrtoormiit, East Greenland (70° 24′ 0′′ N, 21° 58′ 0′′ W). Seals were either shot while hauled out on ice or in the water or caught in nets under the sea ice. Sections of mesenteric lymph nodes were collected from harvested ringed seals within 10– 60 min of death, under Aarhus University's permits and supervision. Sections of lymph nodes were cut into 5 5 mm pieces using a sterile scalpel, and were placed in cold Dulbecco’s modified
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Eagle’s medium (DMEM) culture medium prior to preparing single cell suspensions for cryopreservation in the vapour phase of liquid nitrogen, as previously described for field samples of marine mammals (De Guise et al., 1996). Full depth blubber samples, sections of most major organs, and blood were also collected for contaminant analysis and/or other research programs. Tissues were imported into Denmark and further distributed to send to various collaborators including the University of Connecticut (United States; National Marine Fisheries Service Permit #10081637-03) and the University of Antwerp (Belgium). All tissue samples were stored at 20 °C while lymph node cells were kept in liquid nitrogen until analytical processing. 2.2. Analysis of blubber POP concentrations The analytical protocol was based on previous methods (Weijs et al., 2009) and allowed for the analysis for 42 PCB congeners (CB 18, 28, 31, 44, 47, 49, 52, 66, 70, 74, 87, 95, 99, 101, 105, 110, 118, 128, 132, 138, 146, 149, 151, 153, 156, 158, 170, 171, 172, 174, 177, 180, 183, 187, 194, 195, 196, 199, 203, 205, 206 and 209), several organochlorine pesticides (OCPs), amongst which dichlorodiphenyltrichloroethane (p,p′- and o,p′-DDT) and its metabolites (p,p′- and o,p ′-DDD and p,p′- and o,p′-DDE), hexachlorobenzene (HCB), cis- and trans-chlordanes (CN and TN) and its metabolites oxychlordane (OxC) and cis- and trans-nonachlor (CN and TN), and several PBDE congeners (BDE 28, 47, 49, 99, 100, 153, 154, and 183). An accurately weighed blubber subsample (median: 0.14 g; range: 0.09– 0.23 g) was homogenised and mixed with anhydrous sodium sulphate, spiked with internal standards (CB 143 and BDE 77), and extracted for 2 h with hexane: acetone (3:1, v-v) in an automated hot Soxhlet extractor (Büchi Flawil, CH). The lipid content was determined gravimetrically on 1/10th of the extract (after 1 h at 100 °C) prior to clean up on 8 g acidified silica (44% sulphuric acid). PBDEs, CHLs, and higher PCBs were measured with an Agilent 6890–5973 gas chromatograph coupled to a mass spectrometer system (GC–MS), equipped with a 30 m x 0.25 mm x 0.25 mm DB-5 capillary column (J&W Scientific, Folsom, CA, USA) and operated in electron capture negative ionisation mode. Lower PCBs, DDXs (DDT, DDD, DDE), and HCB were quantified using an identical GC–MS operated in electron ionisation mode and equipped with a 25 m x 0.22 mm x 0.25 mm HT-8 capillary column (SGE, Zulte, Belgium). Procedural blanks, analyzed simultaneously with every batch of seven samples, were consistent (relative standard deviation o 30%) and therefore the mean value was calculated for each compound and subtracted from the sample values. The limit of quantification (MLOQ) was set at 3*standard deviation (SD) of the mean blank measurements, or, for analytes not detected in procedural blanks, set at signal to noise ratio (S: N) ¼10:1. Mean 7SD recoveries of the internal standards CB 143 and BDE 77 were 86 76% and 93 710%, respectively. The analytical procedure was validated through the analysis of certified material SRM 1945 for which deviations from certified values were less than 10%. This quality control scheme has been positively assessed through participation in the NIST/NOAA Intercomparison Programme for Organic and Inorganic Contaminants in Marine Mammal Tissues. 2.3. Analysis of serum PFAS concentrations The extraction method for PFAS in serum was based on solid phase extraction (SPE) as previously described (Keller et al., 2010). Before extraction, the samples (100 ml volume) were spiked with 1 ng of each of the following 13C-labelled internal standards: 13C2PFHxA, 13C4-PFOA, 13C5-PFNA, 13C2-PFDA, 13C2-PFUnA, 13C2-PFDoA, 13 C8-PFOSA, 13C4-PFHxS, 13C4-PFOS, all obtained from Wellington Laboratories (Guelph, ON, Canada). Instrumental analysis was
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performed by liquid chromatography-tandem mass spectrometry (LC-MS-MS) with electrospray ionisation (ESI) in negative mode. The samples were extracted and analyzed in batches together with a procedural blank and two control samples, represented by an aliquot of test material previously analyzed in a ring test and for which assigned values for PFOS and PFOA concentrations are available. The detection limits ranged from 0.02 to 0.7 ng mL 1. As an additional quality control procedure, the performance of analytical method is regularly tested three times a year by participating in the Arctic Monitoring and Assessment Programme (AMAP) Ring Test for Organic Pollutants in Human Serum organized by the Institute Nationale de Santé Publique du Québec. Detection limit and precision of the analytical method were determined using spiked rabbit serum (Sigma Aldrich). 2.4. Preparation of immune cells Cryovials containing lymph node cell suspensions were quickly thawed in a 37 °C water bath and the contents were transferred to a conical tube containing 10 times the volume of warm DMEM supplemented with 1 mM sodium pyruvate, 100 mM non-essential amino acids, 25 mM HEPES, 2 mM L-glutamine, 100 U mL 1 penicillin, 100 mg mL 1 streptomycin and 0.25 mg mL 1 Fungizone (all obtained from Life, Grand Island, NY, USA), along with 10% fetal bovine serum (Hyclone, Logan, UT, USA), thereafter referred to as complete DMEM. Tubes were centrifuged at 220 g for 10 min at room temperature, and the cell pellet was re-suspended in complete DMEM. After thawing and washing once in DMEM, the cell suspension was passed through a sterile pipette containing sterile nylon wool to remove any residual connective tissue. To remove dead cells, the cell suspension was processed using the MACS Dead Cell Removal Kit (Miltenyi Biotec, Auburn, CA, USA), as per manufacturer's instructions. After collecting and washing the live cells, as confirmed by trypan blue staining (viability above 95%), the cell suspension was adjusted to 2 106 cells mL 1. 2.5. Mitogen-induced lymphocyte proliferation Mitogen-induced lymphocyte proliferation, without in vitro chemical exposure, was evaluated by methods routinely performed in our laboratory (Levin et al., 2007a; Mori et al., 2006). Lymphocytes were plated in 96 well flat bottom tissue culture plates, in triplicate, and incubated at 37 °C with 5% CO2 for a total of 66 h with the T cell mitogen concanavalin A (ConA), the B cell mitogen lipopolysaccharide (LPS), or complete DMEM medium alone (no mitogen, unstimulated). Con A and LPS were used at the sub-optimal concentrations of 0.1 mg mL 1 and 0.05 mg mL 1, respectively. Sub-optimal concentrations were used as they proved more sensitive in detecting immunotoxicity (Mori et al., 2006). Lymphocyte proliferation was reported as the stimulation index (SI ¼ mitogen/no mitogen). Mitogen-induced lymphocyte proliferation with in vitro chemical exposure was evaluated as above, though with addition of chemical during incubation. For the PCB congeners, the concentrations chosen (range 0.5–20 mg g 1 ww in culture media) were within effect levels seen in previous in vitro exposure studies (De Guise et al., 1998; Mori et al., 2008; Mori et al., 2006) and at concentrations 2 to 10 fold higher in plasma of ringed seals from the Baltic and Svalbard areas (maximum values of 0.185 and 0.260 mg g 1 ww plasma, respectively) (Routti et al., 2008). When converted to a lipid weight, the ΣPCB concentration range in the Baltic and Svalbard seals were comparable to the ΣPCB concentration range found in the blubber of seals in the current study. For PFOA and PFOS, the concentrations chosen (range 1– 1000 ng g 1 ww) were also within and above the concentrations measured in ringed seal serum in this study (range 0.6–3.1 ng g 1
ww PFOA; range 21.1–196.6 ng g1 ww PFOS). The following chemicals and final concentrations were tested: PCB congeners 138, 153, 180 and 169 (ULTRA Scientific, N. Kingstown, RI, USA) were tested at 0 (unexposed control), 0.5, 1.0, 2.5, 5, 10, 15, and 20 mg g 1. PFOA (Sigma-Aldrich, St. Louis, MO, USA) was tested at 0 (unexposed control), 1, 3, 10, 30, 100, 300, 1000 ng g 1. PFOS (Sigma-Aldrich, St. Louis, MO, USA) was tested at 0 (unexposed control), 1, 3, 10, 30, 100, 300 ng g 1. Both PFASs were tested using the mitogens ConA and LPS, while the PCB congeners were tested with ConA only due to limited lymphocytes available for testing. 2.6. Statistical analysis Since immune cells from the same individual were divided into several chemical exposure groups (in vitro exposure), one-way repeated measures ANOVA with Dunnett's test, using the SigmaStat 3.5 (Jandel Scientific, San Rafael, CA) software, was used to compare the effects at the different in vitro concentrations to unexposed control cells. This was done to determine what concentrations significantly affected proliferation, and determine the lowest concentration that was significantly different from control (LOEC: Lowest Observed Effective Concentration). Concentrationresponse curves were generated by plotting the effects of the different treatment groups as a percentage of the unexposed control cells, and determining mathematically what concentration resulted in a 50% change (decrease or increase) in response (EC50). Correlation analyses using Pearson product-moment correlation were performed using SPSS 16.0 (IBM, New York, USA). For all tests, po 0.05 was used for statistical significance. A t-test was used to compare contaminant concentrations between males/females and adults/subadults.
3. Results Table 1 summarizes the biological and contaminant data for the major POPs (ΣPCBs, ΣDDTs, ΣHCHs, ΣPBDEs, HCB) and PFASs (PFOS and PFOA) for each ringed seal. Full details of all blubber/ serum concentrations are provided in the supplementary information (Table S1). Nine adult or sub-adult seals were captured in total, six males and three females, with all but one (ID 43166) in good body condition. Contaminant concentrations were neither significantly different between males and females nor between adults and sub-adults (p4 0.05 for all contaminants). No significant correlations existed between lymphocyte proliferation (with and without mitogen) and any tissue contaminant concentration in the nine ringed seals (p 40.05 for all comparisons; data not shown). For the in vitro exposure tests, exposure to the non-coplanar CB 138, 180 and 153 significantly reduced ConA stimulated lymphocyte proliferation at 10, 15 and 20 mg g 1 respectively (Fig. 1 and Fig. S1). Furthermore, the relative magnitude of reduction in lymphocyte proliferation was greater for CB 138 than CB 153 and 180. In contrast, the coplanar CB 169 did not significantly modulate ConA stimulated lymphocyte proliferation at any concentration tested (p 40.05; Fig. 1 and Fig. S1). EC50 values were determined at 13.3, 20.7, 20.8, and 54.6 mg g 1 for CB 138, 153, 180, and 169, respectively (Fig. 1). EC50 values 420 were extrapolated beyond the 0–20 range based on the best fit curve. Unstimulated lymphocyte proliferation was not significantly modulated by any PCB congener or concentration tested (p 40.05; data not shown). ConA- and LPS-induced lymphocyte proliferation was not significantly modulated by any concentration of PFOA or PFOS tested (p 40.05; Fig. S2).
M. Levin et al. / Environmental Research 151 (2016) 244–250
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Table 1 Biological and contaminant data for nine East Greenland ringed seals. Serum (ng mL 1)
Blubber (ng g 1 lw)
Seal ID Sex Age
Zoological length (cm)
Blubber thickness (mm)
Weight category
lipid (%) PFOS
PFOA
ΣPCBs
ΣDDXs ΣHCHs ΣPBDEs HCB
43181 43189 43187 43184 43183 43188 43166 43192 43190
122.5 122 110 99 121 104 107 115 110
72 60 40 48 60 45 30 57 40
Very fat Fat Medium Medium Fat Medium Thin Fat Medium
85.4 84.6 88.8 78.5 90.7 81.8 82.7 97.8 92.5
25.6 21.1 31.6 196.6 51.6 49.3 48.2 45.7 46.0
1.1 1.3 0.6 3.1 1.9 1.2 2.6 2.6 1.6
349 349 349.4 349.6 483.5 545 569.2 663 2852.5
264.6 342.2 322.4 273.2 382.9 426.6 521.6 568.9 2761.8
15.7 22.6 28.5 38.6 32.7 37.8 25.1 28.5 51.7
9.1 10.8 10.6 9.6 13.4 14.6 18.8 18.9 49.4
9.2 10.4 10.3 8.9 11.1 6.7 8.6 6.4 12.3
Mean SD Min Max
87 6 78.5 97.8
57.3 53.4 21.1 196.6
1.8 0.8 0.6 3.1
723.4 807.0 349 2852.5
651.6 798.2 264.6 2761.8
31.2 10.5 15.7 51.7
17.2 12.6 9.1 49.4
9.3 1.9 6.4 12.3
F M M F M M M F M
Adult Adult Adult Subadult Adult Subadult Subadult Adult Subadult
125
*
***
***
10
15
20
% control
100
75
50
25
0 0
5
concentration (µg g-1)
Fig. 1. Concentration-response curves for ConA-induced lymphocyte proliferation upon in vitro exposure to increasing concentrations of CB 138, 153, 180, and 169. Data for each congener are expressed as a percent of the unexposed control. EC50 ¼ effective concentration that reduced lymphocyte proliferation by 50%. *, p o0.05 compared to unexposed control for each individual congener (one-way RM ANOVA). Error bars are þ/ 1 SD.
4. Discussion This study utilized blubber, blood, and lymph node immune cells from freshly harvested ringed seal to evaluate the effects of two major classes of environmentally persistent contaminants, i.e. PCBs and PFASs, upon either in vitro or natural exposure, on the immune system. No significant associations were found between mitogen-induced lymphocyte proliferation and tissue contaminant concentrations. However, in vitro exposure to the noncoplanar PCB congeners 138, 153, and 180 significantly reduced lymphocyte proliferation at 10 mg g 1 or above. The lack of significant relationships between lymphocyte proliferation and tissue contaminant concentrations was not surprising, as the concentrations of tissue contaminants were below concentrations shown to modulate immune functions in other marine mammals and laboratory animals. Serum PFOS (range 21–197 ng mL 1 ww) measured in this study was within the range of concentrations previously reported between 1982 and 2010 in East (liver, 20–352 ng mL 1 ww) and West (liver, 12 and 397 ng mL 1 ww) Greenland ringed seals (Riget et al., 2013). Serum PFOA (range 0.6–3 ng mL 1 ww) measured in this study was within the range of concentrations previously reported between 1982 and 2010 in East (liver, below minimum
detection levels to 2.2 ng mL 1 ww) and West (liver, below minimum detection levels to 2.4 ng mL 1 ww) (Riget et al., 2013) Greenland ringed seals. Serum PFOS (mean 57.3 ng mL 1) and serum PFOA (mean 1.8 ng mL 1) measured in the current study were below the lowest observed adverse effect level (LOAEL) in serum that induced immunotoxicity (notably T-cell-dependent IgM antibody response) in laboratory mice (Dewitt et al., 2012), 92 ng mL 1 for PFOS and 74,000 ng mL 1 for PFOA, with one exception, seal ID 43184, which had a serum PFOS concentration of 196.6 ng mL 1. Blubber ΣPCBs (range 349–2852 ng g 1 lw) measured in this study was within the range of concentrations previously reported in East Greenland ringed seals in 2004 (ΣPCB10 mean 605 þ/ 240 ng g 1 lw, ) (Riget et al., 2006) and above the concentration range previously reported for West Greenland ringed seals between 1994 and 2004 (ΣPCB10 133–332 ng g 1) (Vorkamp et al., 2008). TheΣPCB concentrations in this study were below the threshold for physiological effects (1370 ng g 1 lw blubber) as previously reported for aquatic mammals, including ringed seals (Brown et al., 2014; Kannan et al., 2000). Concentrations were, however, within range of the effect threshold for lymphocyte proliferation recently established for marine mammals, as well as for seals specifically. In their meta-analysis of all available marine mammal immunotoxicology studies (including in vitro and in vivo data), Desforges et al. (Desforges et al., 2016) found a sigmoidal dose-dependent decrease in lymphocyte proliferation, with seal threshold effect levels of ΣPCBs of 8 70.01 ng g 1, respectively. Concentrations in this study were also within range of estimated critical body residues (CBR) for immunosuppression in polar bears (Ursus maritimus), 469 ng g 1 lw blubber (Dietz et al., 2015). CBR is defined as the lowest observed total body concentration of a contaminant, which is associated with the occurrence of adverse toxic effects. Although marine mammal CBRs have only been calculated for polar bears, they may serve as a reference for other marine mammals until species-specific CBRs are defined. ΣDDX (range 254–2762 ng g 1) measured in this study was found within a similar concentration range previously reported in East Greenland ringed seals (adult, 610–2460 ng g 1 lw; subadult, 360–1070 ng g 1 lw) (Vorkamp et al., 2004) and the concentration range previously reported in West Greenland ringed seals between 1999 and 2004 (190–549 ng g 1 lw) (Vorkamp et al., 2008). Blubber p,p′-DDT measured in this study was below concentrations found in blubber of harbor seals fed either herring from the ‘uncontaminated’ Atlantic ocean or ‘contaminated’ Baltic Sea (mean 306 and 2448 ng g 1 lw, respectively). Only seals fed the
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Baltic Sea herring had a significant reduction in mitogen-induced T lymphocyte proliferation (de Swart et al., 1996), but this was related to TEQ levels in blubber, and not only DDT specifically. In addition, ΣDDXs in this study (mean 652 ng g 1 lw blubber) was also 10-fold below the estimated CBR for immunosuppression in polar bears, 6502 ng g 1 lw blubber (Dietz et al., 2015). Blubber ΣHCHs (range 16–52 ng g 1 lw) measured in this study was similar to the concentration range previously reported in East Greenland ringed seals (adult, 53–92 ng g 1 lw; subadult, 29–120 ng g 1 lw) (Vorkamp et al., 2004) and similar to the median concentration previously reported in West Greenland ringed seals in 2006 (47 ng g 1 lw) (Vorkamp et al., 2008). Blubber ΣHCHs (mean 31.2 ng g 1 lw) measured in this study was nearly 42-fold below the estimated CBR for immunosuppression in polar bears, 1302 ng g 1 lw blubber (Dietz et al., 2015). Blubber ΣPBDEs (range 9–49 ng g 1 lw) measured in this study was similar to the mean concentration previously reported in East Greenland ringed seals in 2004 (mean 32.5 þ/ 13.8 ng g 1 lw) (Riget et al., 2006) and slightly above the median range previously reported in West Greenland ringed seals (2–8 ng g 1 lw) (Vorkamp et al., 2008). Blubber ΣPBDEs (mean 17.2 ng g 1 lw) measured in this study were below concentrations ( 18 106 ng g 1 lw liver) found to modulate immune functions (i.e. antibody production) in experimentally exposed mink (Martin et al., 2007). Blubber HCB (range 6–12 ng g 1 lw) measured in this study was similar to the concentration range previously reported in East Greenland ringed seals (adult, 9–20 ng g 1 lw; subadult, 8– 27 ng g 1 lw). To date, relationships between HCB and changes in marine mammal immune functions have not been documented (Dupont et al., 2013). Overall, most tissue contaminant concentrations were below concentrations found to induce immunotoxicity in other marine mammals and laboratory animals. The only exception are PCBs where blubber levels appear to fall within range of effect thresholds calculated for marine mammal lymphocyte proliferation (Desforges et al., 2016); thus concentrations in this study may be bordering or just above the no observable effect level (NOEL). A larger sample size is however necessary to observe an in vivo correlation between immune functions and PCB concentrations. To further examine the effects of contaminants on the immune system of ringed seals, in vitro exposure experiments were conducted with four individual PCB congeners, CB 138, 153, 180 and 169, and two PFSAs, PFOA and PFOS. We chose in vitro concentrations, on a weight wet basis (mg g 1 or ng g 1), that were within and above those found upon natural in vivo exposure (in serum or plasma), also on a weight wet basis, in other populations of ringed seals for which published data exist (see Methods). This allowed us to use an environmental relevant concentration range to assess the immunotoxic potential of these chemicals on circulating immune cells using an in vitro approach. In vitro experiments allow for the detection of the direct effects of contaminants on immune functions. When performing the in vitro experiments, cells from the same individual serve as the unexposed control for that individual; therefore, it takes into account any previous POP exposure and other confounding factors such as age and sex. Therefore, any modulation of immune function is the result of the actual in vitro exposure. In addition, Desforges et al. (2016) found that for lymphocyte proliferation, dose response curves for in vitro and in vivo lymphocyte proliferation vs. pollutant concentrations overlapped very well, suggesting that the in vitro approach can be suggestive of the outcome of in vivo exposure to similar concentrations. Importantly, the use of in vitro concentration–response experiments allows the testing of a wider range of concentrations, especially at the higher end of the range, which animals may experience if contaminant burdens continue to
increase. This may significantly improve wildlife risk assessment. In addition, the use of in vitro testing allows one to assess and compare the immunotoxic potential of individual contaminants at the cell level and compare among different species (Levin et al., 2007b, 2007c, 2005b; Mori et al., 2006). In this study, there was a concentration dependent suppression of T lymphocyte proliferation at concentrations of 10 mg g 1 or higher for non-coplanar PCBs. Similar immunomodulatory effects were documented for non-coplanar PCBs in B6C3F1 mouse, managed-care beluga whales, and human T lymphocyte proliferation under comparable in vitro conditions (Leibrecht, 2007). For example, suppressed lymphocyte function was found in mice for PCB congeners 138, 153 and 180 and in beluga for CB 138, however, lymphocyte proliferation increased in beluga exposed to CB 153 and 169 and in humans for all four congeners. Not only is the directionality of in vitro lymphocyte response variable between species, but the effect concentrations and thresholds differ as well. For example, human lymphocytes were most sensitive to CB 138 with LOEC and EC50 values of 5 and 6.9 mg g 1, compared to 10.0 and 13.3 mg g 1 for ringed seals, 15.0 and 19.4 mg g 1 for mouse, and 18.7 and 420 mg g 1 for beluga. Similar to natural exposure, there was no modulation of mitogen-induced lymphocyte proliferation upon in vitro exposure to PFOA or PFOS. For PFOS, the estimated CBR for immunotoxic effects in polar bears was 9356 ng g 1 ww (Dietz et al., 2015), with no value available for PFOA. The LOAEL for PFOS that induced immunotoxicity in laboratory mice, 92 ng mL 1 serum, was within the range of in vitro concentrations tested here, while the LOAEL for PFOA, 74,000 ng mL 1 serum, was above the range tested here (Dewitt et al., 2012).
5. Conclusions Taken together, these data suggest this population of ringed seals is not currently at high risk for immunotoxic effects upon POP or PFAS exposure examined in this study. However, under stressful conditions such as nutritional deficiency and reproduction, marine mammals may undergo periods of blubber metabolism and fat mobilization such that blubber PCBs would concurrently mobilize and reach higher levels in circulation (Debier et al., 2006; Yordy et al., 2010). Thus, the higher concentrations in blood may reach levels demonstrated to temporarily modulate immune functions, potentially increasing an individual's risk for immunomodulation during already stressful periods. In addition, these data provide further ‘weight of evidence’ that PCBs, especially the non-coplanar congeners (CB 138, 153, and 180), are potentially immunotoxic in marine mammals, although difference among species susceptibility must be considered as previously documented (Mori et al., 2006). At current levels, PFASs appear not to be of concern on some aspects of the immune system. Importantly, measuring other aspects of immune functions, such as key innate and adaptive immune functions, including phagocytosis, respiratory burst, and natural killer (NK) cell activity, will provide a broader and clearer picture on the effects of contaminants on the overall health of ringed seals, as well as other top Arctic predators. Ideally, similar studies as the present should be conducted for marine mammals at risk including seals from the Baltic seals and other marine mammals foraging at higher trophic levels, including e.g. harp seals (Pagophilus groenlandicus), hooded seals (Cystophora cristata), toothed whales like narwhals (Monodon monoceros), beluga whales, pilot whales (Globicephala melas) and killer whales (Orcinus orca) as well as polar bears (Ursus maritimus).
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Funding sources Major funding for this project was provided by the DANCEA (Danish Cooperation for Environment in the Arctic) and ARC (Arctic Research Centre). Additional travel support (for Levin) was provided by the University of Connecticut Research Foundation and the University of Connecticut's Interdisciplinary Research and Training Initiative on Coastal Ecosystems and Human Heath (IRICH). Funding travel support was also provided by the Norwegian University of Science and Technology, N-7491, Trondheim, Norway.
Acknowledgements We thank local subsistence Scoresby Sound hunters for helping with the sample collection at Kap Tobin in February–March 2012, as well as colleagues from four countries (Denmark, Norway, Canada, United States) as part of the AURORAE project. Major funding for this project was provided by the DANCEA (Danish Cooperation for Environment in the Arctic) and ARC (Arctic Research Centre). Additional travel support (for Levin) was provided by the University of Connecticut Research Foundation and the University of Connecticut's Interdisciplinary Research and Training Initiative on Coastal Ecosystems and Human Heath (I-RICH). Funding travel support was also provided by the Norwegian University of Science and Technology, N-7491, Trondheim, Norway.
Appendix A. Supplementary material Supplementary data associated with this article can be found in the online version at http://dx.doi.org/10.1016/j.envres.2016.07. 013.
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