Ecotoxicology and Environmental Safety 113 (2015) 248–258

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Comparison of produced water toxicity to Arctic and temperate species L. Camus a,f, S. Brooks c, P. Geraudie a, M. Hjorth d, J. Nahrgang a,b, G.H. Olsen a,n, M.G.D. Smit e a

Akvaplan-niva, High North Research Centre, 9296 Tromsø, Norway UiT—the Arctic University of Norway, Faculty of Biosciences, Fisheries and Economics, Department of Arctic and Marine Biosciences, NO-9037 Tromsø, Norway c Norwegian Institute for Water Research (NIVA), NO-0349 Oslo, Norway d COWI, Parallelvej 2, 2800 Kongens Lyngby, Denmark e Statoil ASA, Rotvoll, 7005 Trondheim, Norway f UiT— the Arctic University of Norway, Faculty of Science and Technology, Department of Engineering and Safety, NO-9037 Tromsø, Norway b

art ic l e i nf o

a b s t r a c t

Article history: Received 8 July 2014 Received in revised form 30 November 2014 Accepted 2 December 2014

Produced water is the main discharge stream from oil and gas production. For offshore activities this water is usually discharged to the marine environment. Produced water contains traces of hydrocarbons such as polycyclic aromatic hydrocarbons as well as alkylphenols, which are relatively resistant to biodegradation and have been reported to cause adverse effects to marine organisms in laboratory studies. For management of produced water, risk-based tools have been developed using toxicity data for mainly non-Arctic species. Reliable risk assessment approaches for Arctic environments are requested to manage potential impacts of produced water associated with increased oil and gas activities in Arctic regions. In order to assess the applicability of existing risk tools for Arctic areas, basic knowledge on the sensitivity of Arctic species has to be developed. In the present study, acute and chronic toxicity of artificial produced water for 6 Arctic and 6 temperate species was experimentally tested and evaluated. The hazardous concentrations affecting 5% and 50% of the species were calculated from species sensitivity distribution curves. Hazardous concentrations were compared to elucidate whether temperate toxicity data used in risk assessment are sufficiently representative for Arctic species. From the study it can be concluded that hazardous concentration derived from individual species' toxicity data of temperate and Arctic species are comparable. However, the manner in which Arctic and non-Arctic populations and communities respond to exposure levels above established thresholds remains to be investigated. Hence, responses at higher levels of biological organization should be studied to reveal potential differences in sensitivities to produced water between Arctic and non-Arctic ecosystems. & 2014 Elsevier Inc. All rights reserved.

Keywords: Arctic Produced water Species sensitivity distribution curves (SSDs) Environmental risk assessment

1. Introduction Oil and gas activities are increasing in the high north. To ensure protection of the environment in new areas of development, risk assessment tools are used to characterize and quantify environmental risk. In today's risk assessment procedures applied to the marine environment, individual species tolerance to chemical exposures are used to allow operators to characterize their potential environmental impacts (Chapman, 1995; Forbes and Calow, 2002). However, these tolerance data are mainly developed for non-Arctic species. As the oil industry is moving into the Arctic, questions arise about the applicability of existing environmental risk assessment tools for produced water (PW) management in the n

Corresponding author. E-mail address: [email protected] (G.H. Olsen).

http://dx.doi.org/10.1016/j.ecoenv.2014.12.007 0147-6513/& 2014 Elsevier Inc. All rights reserved.

Arctic. One of the main uncertainties is the sensitivity of Arctic organisms to PW compared to non-Arctic species. Basic knowledge on the sensitivity of Arctic species has to be in place before established risk assessment procedures and tools can be applied to the Arctic with a sufficient level of certainty. If required risk assessment procedures need to be adapted to Arctic environments. Production of oil and gas generates large volumes of PW. The total volume of PW discharged was 131 million m3 on the Norwegian continental shelf (NCS) in 2012 (Norwegian Oil and Gas, 2013). PW is a complex mixture of formation water (water trapped for millions of years in a geologic reservoir), condensation water and occasionally injection water, injected in the well to maintain production levels. It contains numerous dissolved and particulate organic and inorganic substances with a concentration largely depending on reservoir characteristics. These substances include inorganic salts, metals, radioisotopes and organic compounds, such as polycyclic aromatic hydrocarbons (PAHs), alkylphenols

L. Camus et al. / Ecotoxicology and Environmental Safety 113 (2015) 248–258

(APs) and organic acids (Durell et al., 2006; Johnsen et al., 2004, Neff et al., 2011, Utvik, 1999). Although overall long term effects of PW discharges to the marine environment are likely to be small (Geraudie et al., 2014), many of the substances present in PW have been reported to possess the ability to cause adverse effects to marine organisms (Tollefsen et al., 1998; Meier et al., 2002; Sundt et al., 2009a, 2009b). The large discharge volumes, the complex chemical composition and the lack of knowledge on possible long term ecological impacts has made PW discharges a strong target for concern and research in recent years. In order to assess the potential effects of PW discharges on the Norwegian Continental Shelf, the Water Column Monitoring Programme has been undertaken by the operators since 1997. The monitoring programme has been conducted at various fields within the Norwegian sector (Bakke et al., 2011; Brooks et al., 2009; Hylland et al., 2008; Nilssen and Bakke, 2011; Sundt et al., 2011) and has, for instance, demonstrated the environmental benefits of improved PW treatment system at the Ekofisk field (Brooks et al., 2009). Beyer et al. (2012) performed an environmental risk assessment (ERA) for chosen areas in the North Sea and concluded that environmental exposure of fish to Alkylphenols (APs) from PW is most likely too low to affect reproduction in wild populations of fish in the North Sea. However, a review of recent research on the biological effects of PW and drill cuttings with focus on the Norwegian Continental Shelf suggests concerns linked to effects of produced water (Bakke et al., 2013). AP and polyaromatic hydrocarbons (PAH) from produced water accumulate in cod and blue mussel caged near outlets, but are rapidly metabolized in cod. APs, naphtenic acids, and PAHs may disturb reproductive functions, and affect several chemical, biochemical and genetic biomarkers. Toxic concentrations seem restricted to o2 km distance. The risk of widespread, long term impact from the operational discharges on populations and the ecosystem is presently considered low, but this cannot be verified from the published literature. Authorities and industry need to seek regulatory acceptance of a planned chemical discharge into the environment. In order to manage the environmental risk from PW discharges a dedicated tool has been developed by the oil industry. This risk-based management tool, called the Environmental Impact Factor (EIF), identifies the potentially most harmful substances in PW which enables taking further steps in selecting the most optimal and cost efficient mitigation measure to reduce harm to the environment (Johnsen et al., 2000; Smit et al., 2011). In the EIF methodology, that follows the generic concept for environmental risk assessment (ERA) as described by the United States' Environmental Protection Agency (USEPA, 1993) and the European Commission (EC, 2003), generic toxicity data is used to establish environmental thresholds (e.g. Predicted No Effect Concentrations (PNECs)). In 2012 the Oslo–Paris Convention (OSPAR) adopted a recommendation on a risk-based approach to the management of PW discharges (OSPAR, 2012a) from offshore installations and OSPAR Guidelines were also established (OSPAR, 2012b). The OSPAR convention guides international cooperation on the protection of the marine environment of the Northeast Atlantic (www.ospar. org). The OSPAR framework follows EU-ERA principles (ECHA, 2008) and may elect each OSPAR country to use a substance based approach (e.g. EIF) or a whole effluent approach, or a combination of these approaches. For the purpose of the substance based risk approach, a harmonized list of PNEC values were established for naturally occurring substances in PW, that was based on already established and publicly available values. This list of PNECs was recently implemented for use in EIF PW. An important question that needs answering is whether the toxicity information used in the EIF methodology is transferrable to the Arctic. A number of biological and physical differences

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between Arctic and temperate systems exists (AMAP, 1998). E.g., differences in temperature between Arctic and temperate systems may alter the physical behaviour of oil. In the Arctic ecosystem, the exposure of biota to volatile fraction of hydrocarbons may be prolonged due to lower evaporation rates (MacDonald et al., 2008). Responses of Arctic and non-Arctic organisms may also be influenced by differences in physiological characteristics, including metabolism, lipid fraction, antioxidant levels, and resistance to freezing (Chapman and Riddle, 2003). In recent years, several studies have compared oil related toxicity data for Arctic and non-Arctic species (Skadsheim et al., 2009, Bechmann et al., 2010; Hansen et al., 2011; Jonsson and Björkblom 2011; Olsen et al., 2011, De Hoop et al., 2011, Gardiner et al., 2013). Generally, the differences in sensitivity between Arctic and temperate species in these studies are small. The studies that were performed with 2-methylnaphthalene (Olsen et al., 2011), crude oil (De Hoop et al., 2011) and dispersed oil Gardiner et al (2013) showed that the difference in sensitivity between Arctic and non-Arctic species is insignificant. However, no information is available on the potential effects of PW mixtures on Arctic species and how PW effect levels for Arctic species compare to effect levels for non-Arctic species. In the present study acute and chronic toxicity tests were performed with artificial PW to derive effect levels for 5 standard temperate test species and 6 Arctic species, representing 5 taxonomic groups. The experimental results were used to quantify the concentration causing effects to 50% of exposed individuals (EC50) and no-effect concentration (NEC). These values were used to develop an Arctic and a temperate species sensitivity distribution curves (SSDs) for artificial PW. The hazardous concentrations (the concentration of chemical that is harmful to 5% of species (HC5) and 50% of species (HC50)) were derived from the SSD curves and compared. Based on these results, we discuss the applicability of temperate toxicity data for environmental risk assessment purposes for produced water in Arctic areas.

2. Materials and methods 2.1. Ethical statement All work were performed according to and within the regulations enforced by the Norwegian Animal welfare authorities. The experiments were approved and performed by trained personal with appropriate certificates (FELASA C) for animal experimentation. The experiments were following the OECD guidelines for acute toxicity testing. 2.2. Artificial produced water media The exposure media, described as artificial produced water (PW), was prepared in the laboratory by combining the major classes of PAHs and APs at concentrations typically found in PW from offshore oil and gas platforms in the North Sea (based on Holth et al., 2008). Propionic acid was added to the mixture as the surrogate organic acid fraction at a concentration based on the mean total organic acid concentration discharged from offshore platforms in the North Sea during 2009 (OLF, 2010). The actual concentrations of the PAHs, APs and organic acids in the artificial PW are presented in Table 1. A concentration of 1x PW was considered to be equivalent to the median concentration of PW substances at the outlet of the discharge pipe, prior to dilution with the receiving waters, for a typical offshore oil and gas platform in the North Sea (Holth et al., 2008). For all bioassays, a concentrated working stock solution of the PW mixture was dissolved in acetone before being diluted with

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Table 1 The concentration of compounds used for 1x PW concentration (corresponding to fraction 1) equivalent to the median concentration of chemicals in the PW as it leaves the discharge outlet of a typical North Sea Oil and Gas platform (from Holth et al., 2008). Compound

Concentration

Compound

Concentration

Naphthalene Acenaphthene Fluorene

310 1.7 12.4

340 188.7 112.6

Anthracene Phenanthrene Dibenzothiophene Pyrene Phenol

0.7 16 3.8 0.7 1400

p-Cresol 4-Ethylphenol

1500 600

4-n-Propylphenol 4-Tert-butylphenol

100 40

2-Methylnaphthalene 2,6-Dimethylnaphthalene 2,3,6Trimethylnaphthalene 4-Methyldibenzothiophene 4-Ethyldibenzothiophene 1-Methylphenanthrene 2-Methylanthracene 1,5/1,7Dimethylphenanthrene 9,10-Dimethylanthracene 1,2,6Trimethylphenanthrene Propionic acid pH adjusted with NaOH

4.6 6.6 21.6 2 23.7 2 3.5 201,520

filtered (0.45 mm) seawater to obtain the appropriate range of PW test concentrations for each bioassay. The concentration of solvent in each test media was minimal and did not exceed the test acceptability criteria of 0.1 ml/L. A solvent control at 0.1 ml/L was used in all tests. The artificial PW was pH adjusted with sodium hydroxide to achieve acceptability criteria for each bioassay. Water samples were taken from the bioassays and frozen at
20 °C for chemical analysis. Chemical analyses were carried out by an accredited laboratory (Intertek West Lab, Norway). Toxicity data were expressed based on the measured water chemical concentrations. Average, minimum and maximum values of the ratio between nominal and measured concentrations were calculated for tests with Arctic and non-Arctic species. For some exposures where it was not possible to measure the exposure concentrations (e.g. due to the low exposure volumes), the actual exposure concentration was then assessed based on the average ratios between the nominal and measured concentrations obtained for similar exposures. 2.3. Selection of test species All selected temperate species are routinely used for standard toxicity tests under the framework of the 1992 Oslo–Paris Convention (OSPAR). This convention guides international cooperation on the protection of the marine environment of the Northeast

Atlantic (www.ospar.org). The species were kept either in culture at the Norwegian Institute for Water research (NIVA) in Oslo where the tests were performed, or purchased from laboratories that culture the species. The temperate test set included an acute algal growth test with Skeletonema costatum, an acute survival and a chronic reproduction test with the copepod Tisbe battagliai, an oyster embryo development test with Crassostrea gigas, a brown shrimp acute survival test with Crangon crangon and finally, an acute fish survival test using Scophthalmus maximus. The Arctic test set included an acute algal growth test with Porosira glacialis, an acute survival test with the copepod Calanus glacialis, a bivalve larvae malformation test with Mytilus edulis, a shrimp acute survival test with Sclerocrangon boreas and finally an acute fish survival test using Boreogadus saida and Leptoclinus maculatus. A chronic reproduction test with C. glacialis was also performed, but due to low reproduction in the controls and exposures of this test the results could not be used in the analysis described in this paper. The Arctic species were tested using generally the same test protocols as the standard test species, with minor adjustments. Both the Arctic and temperate species were encompassing 5 different taxonomic groups (Table 2), to ensure that representatives of different taxonomic and functional groups are represented in the study. The toxicity tests with Arctic species were adjusted for lower temperatures. The main deviations from the original protocols concerned the microalgae P. glacialis (starting cell density and minimum growth rate) and the M. edulis embryo assay (development time to trocophore and length of exposure). A general description of the protocols applied is provided in the following section. Details of the protocols used are described in Nahrgang et al. (2011). 2.4. Acute tests 2.4.1. Algal growth test 2.4.1.1. Temperate algae S. costatum. The unicellular algae S. costatum have a size ranging from 2 to 21 mm. The effects of the artificial PW mixture on the growth of S. costatum were studied in accordance with the International Organization for Standardisation (ISO) 10253 guideline (ISO, 2006). The growth medium was prepared by adding ISO 10253 stock solutions (ISO, 2006) to 0.45 mm filtered seawater collected from a depth of 60 m from the outer Oslo fjord, Norway. The growth of the algal inoculum (5  103 cells/L) placed on an orbital shaker in continuous cool white fluorescent light (687 4 mmol m
2 s
1, Philips TLD 36W/950) under constant temperature

Table 2 Acute and chronic toxicity data (expressed as fraction of the undiluted produced water) measured for the artificial produced water with Arctic and temperate species. The effect endpoint used in the acute tests were growth (algae) and survival (other species). For the chronic tests these were development (M. edulis) and reproduction (T. battagliai). Species

Taxonomic group

Region

EC50

st. dev.

EC50 at time

NEC

st. dev

Acute tests Porosira glacialis Boreogadus saida Leptoclinus maculatus Sclerocrangon spp. Calanus glacialis Crangon crangon Crassostrea gigas Skeletonema costatum Tisbe battagliai Scophtalus maximus

Algae Fish Fish Shrimp Copepod Shrimp Mollusc Algae Copepod Fish

Arctic Arctic Arctic Arctic Arctic Temperate Temperate Temperate Temperate Temperate

0.36 0.85 0.87 1.05 5.25 0.38 0.46 0.63 1.69 1.87

0.11 0.01 0.01 1.83 2.20 0.07 0.09 0.02 0.17 –

72 h 96 h 96 h 96 h 96 h 96 h 24 h 72 h 96 h 96 h

n.d. 0.81 0.65 1.05 0.23 0.38 n.d. 0.61 1.18 1.87

n.d. 0.02 0.01 1.84 0.14 0.28 n.d. 0.09 0.14 n.d.

Chronic tests Mytilus edulis Tisbe battagliai

Mollusc Copepod

Arctic Temperate

0.30 0.18

1.71E þ4 0.001

12 d 14 d

0.28 0.11

6.81E þ6 0.02

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(20 71 °C) were measured every 24 h for the duration of the test (72 72 h) using a Beckman Coulter Multisizer 3 (Beckman Coulter, USA). Three replicates were used per treatment and six replicates for the control. An additional solvent (acetone) control was used at 0.1 ml/L. The treatments included 5.6, 3.2, 1.8, 1.0, 0.56, 0.32, 0.18, and 0.1x PW. The average growth rate for each test concentration was calculated using the equation: μ = ln(Nn) − ln(N0)/tn − t0 where Nn is the cell density at time tn and N0 is the cell density at time zero (t0). The percentage inhibition of growth rate as compared to the control was calculated for each treatment. 2.4.1.2. Arctic algae Porosira glacialis. P. glacialis (Grunow) is a cold water marine alga with a size of 30–40 mm, which is found in the Arctic (Martin et al., 2001). It is pelagic, but it can also be found attached to substrate (i.e., pack ice). It is larger than S. costatum and cannot grow at similar cell densities as S. costatum. The Arctic alga P. glacialis was the only cold water species available from cultures (held at the UiT—The Arctic University of Norway, Norway). Three replicates were used per treatment and six replicates for the controls and acetone controls. The treatments included 10, 4.5, 2.1, 0.9, and 0.4x PW. The inoculum size and the maximum growth rate were reduced compared to the original Standard Operating Procedure (SOP) to maintain a healthy culture. The cultures are available through the whole year and the adjustments made in terms of temperature (5 °C), cell density (0.5  106 cells/L) and growth rate (0.3 d
1) were successful compared to the original SOP (20 °C, 5  106 cells/L, 1.8 d
1).

2.4.2. Copepod survival test 2.4.2.1. Temperate copepod T. battagliai. Acute mortality of the marine copepod T. battagliai exposed to the artificial PW mixture was studied in accordance with the ISO 14669 guideline. Copepod survival was observed following exposure to a range of artificial PW concentrations. The treatments included 5.6, 3.2, 1.8, 1.0, 0.56, 0.32, 0.18, and 0.1x PW in addition to an aged seawater control and a solvent control (0.1 ml/L acetone). The copepods were exposed for 487 2 h at 20 71 °C at a pH and dissolved oxygen concentration range of 7.96–8.06 and 6.8–7.4 mg/L respectively, with 20 individuals per treatment. Mortalities were recorded after 24 and 48 h of exposure. Copepods were considered dead if they failed to exhibit any swimming or appendage movement within 10 s after gentle agitation of the test vessel or prodding with a mounted needle. 2.4.2.2. Arctic copepod C. glacialis. C. glacialis (Jaschnov, 1955) is a highly abundant copepod in Arctic waters during spring time (Hirche et al., 1991; Niehoff et al., 2002; Madsen et al., 2008). Their life cycle appears to be tightly coupled to the pack ice ecosystem that provides sea ice algae which the copepod grazes upon in spring time. The copepod is an important source of food to many fish (Kaardtvedt, 2000), seabirds (Karnovsky et al., 2008, 2010) and marine mammals (Laidre et al., 2007). Copepods were sampled in May 2011, in Disko Bay near Qeqertarsuaq, Greenland approximately 1 nautical mile off the coast onboard RV Porsild (Arctic Station, Copenhagen University). A Niskin 30 L water bottle was used to get water from 100 m, which was filtered through a 0.2-mm filter and used in the laboratory exposure experiments. Females of C. glacialis were sampled from plankton tows with a WP-2 nylon net (200 mm mesh with a 1-L non-filtering cod-end) from 250 m to the surface. After collection, the content of the cod-end was immediately diluted with surface seawater and kept dark in a thermo box at approximately 0 °C, while being transported back to the laboratory. Subsequent sorting

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of individuals for the exposure experiments was done under a dissecting microscope in ice-chilled conditions. The 96 h exposure experiment was initiated on collected copepods approximately 48 h after the collection (Nahrgang et al., 2011). The experimental design consisted of 7 treatments with 4 replicates in 1 L glass bottles containing 0.2-mm filtered seawater with pH ranging from 8.16 to 7.76, dissolved oxygen levels were maintained in the interval 6.82–7.43 mg/L, and a salinity of 34– 35 psu, and temperatures in the range of 4.5–6 °C. The experiment was carried out in a laboratory under dimmed light conditions. The copepods were not fed during the experiment. Exposure water was exchanged and live and immobile animals were counted in 24 h intervals. The procedure included gentle filtration of the animals on 200 mm filters, transfer to glass petri-dishes for observation under a dissecting microscope and transfer back to identical glass bottles with new batches of artificial PW. The whole procedure took a maximum of 5 min for one replicate bottle and was performed in ice-chilled conditions. Artificial PW was prepared daily by diluting the same concentrated stock solution with filtered seawater in quantities of 4 L of every treatment solution in Erlenmeyer flasks on a magnetic stirrer to ensure homogenous dissolution. The concentrations range used was 0.21, 0.47, 1.03, 2.27 and 5x PW. The solvent carrier was acetone, and consequently filtered (0.45 mm) aged seawater and acetone at (0.1 ml/L) were chosen as control treatments. The artificial PW was kept in closed containers after mixing to avoid loss of volatile compounds. 2.4.3. Shrimp survival test 2.4.3.1. Temperate brown shrimp C. crangon. Effects of the artificial PW on the survival of the brown shrimp C. crangon was adapted from the OECD 203 test guidelines for acute fish toxicity. Brown shrimp were collected from the sub-tidal zone near to the NIVA marine research station in the outer Oslo fjord, Norway. The shrimps were kept in glass aquaria within a flow-through seawater supply for one week prior the exposure. For the test, shrimps were randomly distributed between treatment tanks (15 L glass aquaria) with 10 individuals per treatment. The concentrations used were 5.6, 3.2, 1.8, 1.0 and 0.56x PW in addition to an aged seawater control and a solvent control (0.1 ml/L acetone), with 50% of each test media replenished after 48 h. The shrimp were exposed for 9672 h at 1571 °C at a pH and dissolved oxygen concentration range of 7.48–8.06 and 7.4–7.8 mg/L respectively. Mortality was recorded every 24 h throughout the duration of the test (Nahrgang et al., 2011). 2.4.3.2. Arctic shrimp S. boreas. S. boreas (Phips) is a decapod which belongs to the Crangonidae family as its temperate counterpart species C. crangon. It is a boreal, widely distributed species in the Arctic and was sampled with an Agassiz dredge deployed from RV Helmer Hansen at Hollenderbukta in Isfjorden, in the Svalbard archipelago. Exposure temperatures were 6.0–6.5. The exposure of S. boreas to artificial PW was adapted from the 96 h assay of C. crangon with the following modifications. The exposure was run in 20 L polyethylene tanks containing 10 animals. S. boreas was exposed to artificial PW at concentration of 10, 4.5, 2.1, 0.9, 0.4x PW including a control and solvent control. The exposure was performed during 96 h at 6.0–6.5 °C in total darkness. 2.4.4. Fish survival test 2.4.4.1. Temperate fish S. maximus. Effects of the artificial PW on the survival of juvenile turbot, S. maximus was based on the OECD 203 test guideline. Juvenile turbot were transported to the NIVA marine research station near Oslo, Norway from a turbot hatchery (Maximus AS, Denmark). The turbot fry (5–10 g) were held in large 200 L aquaria with a flow through seawater supply for 2–3 weeks prior to the experiment, and were fed with commercial fish food

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once daily. The turbot were randomly distributed into 15 L glass aquaria with 7 individuals per treatment. The treatments included 5.6, 3.2, 1.8, 1.0 and 0.56x PW in addition to an aged seawater control and a solvent control (0.1 ml/L acetone), with 50% of each test media replenished after 48 h. The turbot were exposed for 967 2 h at 15 71 °C at a pH and dissolved oxygen concentration range of 7.77–8.11 and 7.0–7.9 mg/L respectively. Mortality was recorded every 24 h throughout the duration of the test (Nahrgang et al., 2011). 2.4.4.2. Arctic juvenile fish B. saida. The polar cod, B. saida, is a small gadoid fish (Bakke and Johansen, 2005) with a life span of no more than 7 years (Bradstreet et al., 1986). Polar cod has a circumpolar distribution and has been found as far north as the North Pole (Andriyaschev, 1954). The fish were sampled onboard RV Helmer Hanssen (UiT) using a fish lift system (Holst and McDonald, 2000) which reduces stress on the animal during the catch. Fish were then kept in quarantine until no mortality was observed. The exact same protocol was used for B. saida as for S. maximus.

formed shell and contained meat. Small deformation of the shell, e.g., pinch to the hinge, was still considered as normal as long as it was still fully D-shaped. 2.5.2.2. Arctic bivalve M. edulis. Adult M. edulis were obtained from a local mussel farm (Lyngen Alps, Northern Norway). Single individuals were placed in separate vials. Gametes were obtained by a process of thermal shock (and prozac) and in vitro fertilized. Larvae (25 per ml) were placed in each well (4.5 ml) of a 12 wells plate 4 replicates per concentrations at the temperature of 4 °C and exposed from the trocophore stage to PW concentrations (10, 4.5, 2.1, 0.9, 0.4, 0.19, and 0.08x PW), a control and solvent control, until the development of the D-shape. Although originally the bivalve development test is an acute test (24 h for oyster larvae), in this case the results can be considered as chronic responses. Due to the slower development rate of the M. edulis larvae under Arctic conditions the exposure was conducted during 12 days instead of 24 h in order to reach the endpoint of development (Nahrgang et al., 2011). 2.6. Data analyses

2.4.4.3. Arctic juvenile fish L. maculatus. L. maculatus is found in the northern hemisphere, distributed from Arctic to temperate waters. It inhabits sandy bottoms and muddy bottoms, feeding on polychaetes and crustaceans (Eschmeyer et al., 1983). The juvenile fish were sampled onboard RV Helmer Hanssen (UiT) with a plankton net during night (the juvenile fish were migrating up and down the water column). Fish were then kept in quarantine until no mortality was observed. The exact same protocol was used for L. maculatus as for B. saida and S. maximus. 2.5. Chronic tests 2.5.1. Copepod reproduction test 2.5.1.1. Temperate copepod T. battagliai. The effects of the artificial PW mixture on the long term survival and reproductive success of the marine copepod T. battagliai were studied in a 14 d exposure, based on the work performed by Hutchinson et al. (2009). Ten copepods in separate test chambers (wells of a 12 well plate), containing 5 ml of test solution, were used per treatment. Ovigerous females were randomly added to each test vessel (Po generation). The treatments included 10, 5.6, 3.2, 1.8, 1.0, 0.56, 0.32, 0.18, and 0.1x PW concentration in addition to an aged seawater control and a solvent control (0.1 ml/L acetone). Water test temperatures were 20.6–21.5 °C. The test solutions were renewed every 2 or 3 days, at this time the surviving Po generation females were transferred to new chambers of appropriate test media. The animals were fed with a concentrated algal stock (Rhodomonas reticulate 2  105 cells/mL) during each renewal. Observations were made daily for mortality (absence of movement) and for the presence and number of offspring (F1 generation) in each test chamber. 2.5.2. Bivalve developmental test 2.5.2.1. Temperate bivalve C. gigas. Effects of the artificial PW on the development of embryos of the oyster, C. gigas, were studied in accordance with the ASTM E724-89 (1993) guidelines. In brief, oyster trocophore larvae (2 h post fertilization), incubated at 2471 °C were exposed to a range of PW concentrations for 2472 h. During this time the larvae developed into the D-shaped veliger larvae. The effect of PW chemical exposure (5.6, 3.2, 1.8, 1.0, 0.56, 0.32, 0.18, and 0.1x PW) on D-larval development was assessed microscopically (200x magnification) after the exposure period. The embryo density within the test media was approximately 50 individuals per ml, with 4 replicates per treatment and controls. Normal D-larvae were considered to possess a completely

2.6.1. Effect concentrations (EC50) and no effect concentrations (NEC) From all observations median EC50 and NEC were calculated using a Dynamic Energy Budget model for toxicity (DEBtox model) (Kooijman and Bedaux, 1996). The EC50 represents the median effect concentration affecting 50% of the exposed population after a specific exposure time. The NEC represents the concentration below which effects never will occur, even not at chronic exposure times. The DEBtox model is based on the dynamic energy budget theory that embodies a set of simple rules for how organisms acquire resources, and subsequently use these resources to fuel metabolic processes such as growth, development and reproduction. For detailed explanations of how DEBtox calculates NEC and LC50 values the reader is referred to Kooijman and Bedaux (1996). The calculations were based on the fraction of artificial PW components in the exposures. In order to calculate the artificial PW fraction the sum of the measured concentrations of all components (sum total hydrocarbons) at the start of each experiment was compared with the sum of the components originally present in the artificial PW profile. For some of the tests only in the high, low and medium exposures concentrations were measured. The measured concentrations in the other exposures in the same test were assessed based on the average ratio between the nominal and the measured concentration from the low, medium and high exposures in the same test. Because of the low exposure volumes in the tests with M. edulis and P. glacialis for these tests measured concentration was assessed based on the average nominal/measured concentration ratio for Arctic tests. 2.6.2. Species sensitivity distribution curves (SSDs), 5% hazardous concentration (HC5) and 50% hazardous concentration (HC50) In the SSD approach, single-species test data are combined to predict concentrations affecting a defined percentage of species in a community (Kooijman,1987), known as the potential affected fraction of species (PAF). This is accomplished by fitting LC50 or NEC values versus chemical concentration using a distribution such as the log-normal or log-logistic equation. The distribution derived by this curve fitting procedure is known as a species sensitivity distribution (SSD). Hazardous concentrations, the concentration of chemical that is harmful to a defined percentage of species, typically reported for 5% of species (HC5) and 50% of species (HC50), are then derived from the constructed SSD curve. HC5 and HC50 are used in risk assessments to establish acceptable environmental practices (Van Straalen and Denneman, 1989; Aldenberg et al., 2002).

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Arctic and temperate SSDs were constructed from NEC and EC50 values using a log-normal distribution (Aldenberg et al., 2002; Posthuma et al., 2002; Gaddum, 1945). Each log transformed toxicity data series was tested using the Anderson–Darling goodness-of-fit test to verify that the data were normally distributed. The Anderson–Darling test highlights differences between the input data and the tail of the derived log-normal distribution and is regarded as one of the most powerful statistical tests for detecting departures from normality (Aldenberg et al., 2002). Lower, median and upper estimates for the 50% hazardous concentration (HC50) and the 5% hazardous concentration (HC5) were calculated according to the procedure described by Aldenberg and Jaworska (2000). A t-test and an F-test were applied to test for significant differences in respectively the means and the variances of the Arctic and temperate SSDs.

3. Results 3.1. EC50 and NEC The chemical analyses showed that on average the measured fractions of PW at t ¼0 were 32% of the nominal concentrations for the Arctic exposures with a minimum of 23% for the test with S. boreas and a maximum of 36% for the acute test with C. glacialis. For the temperate exposures the average was 69% with a minimum of 52% for S. costatum and a maximum of 80% for S. maximus. These results indicate lower solubility of PW components at lower temperatures. Arctic and non-Arctic EC50 values ranged from 0.30 to 5.25 and 0.18 to 1.87 (fraction artificial PW) respectively, while Arctic and non-Arctic NECs range from 0.23 to 1.05 and 0.11 to 1.87 (Table 2). The chronic tests with the temperate copepod T. battagliai and the Arctic bivalve M. edulis showed the lowest EC50 values. Among the Arctic species, at the EC50 level, the algae were the most sensitive while the copepod appeared to be the most robust. For the temperate species and, again based on EC50s, the algae were also among the sensitive species together with the shrimp and the oyster embryo with the shrimp being the most sensitive. The EC50 for the Arctic shrimp was almost 3 times higher than the temperate shrimp and showed sensitivity in the same range as the Arctic fishes. The EC50 of the temperate fish was over two times higher than the Arctic fish species with an EC50 value in the same range as the temperate copepod. DEBtox was not able to estimate NEC values for all species due to the fact that for some tests the data did not fully cover the complete concentration–effect–time relationship. Of the species for which a NEC was established, the copepod seemed the most sensitive Arctic species with a NEC comparable to the NEC calculated for M. edulis development. For the temperate species the copepod reproduction was the most sensitive endpoint while the temperate fish still remains the least sensitive species based on the NEC. 3.2. SSD, HC5 and HC50 Fig. 1 presents the SSD curves for Arctic and temperate species based on the acute EC50 values. As a reference also the chronic EC50 values for M. edulis and T. battagliai are shown in this figure, but these two values have not been used to construct the SSD. Fig. 2 presents the SSD curves for Arctic and temperate species based on the chronic NEC values derived from both the acute and chronic tests. Since for T. battagliai, two chronic NEC values are available (one from the survival test and one from the reproduction test, only the most sensitive endpoint was used to construct the SSD (reproduction)).

Fig. 1. Species sensitivity distributions based on acute EC50 values for 5 Arctic species (solid line) and 5 temperate species temperate species (dashed line). For reasons of comparison the two chronic EC50 values for M. edulis larvae development and T. battagliai reproduction have also been plotted in this graph, but these values have not been used to construct the SSD.

Fig. 2. Species sensitivity distributions based on chronic NEC values for 5 Arctic species (solid line) and 5 temperate species (dashed line). Chronic NEC values derived from both acute and chronic tests have been used to construct the SSD.

Table 3 Upper, median and lower estimates of the HC5 and HC50 values derived from the species sensitivity distributions based on EC50s (Fig. 1).

Arctic species Temperate species

HC5 median estimate

Lower–upper estimate

HC50 median estimate

Lower–upper estimate

0.19 0.22

(0.02–0.48) (0.04–0.44)

1.08 0.81

(0.42–2.74) (0.40–1.64)

Table 3 presents the upper, median and lower estimates of the HC5 and HC50 values based on the acute EC50 values while Table 4 presents the upper, median and lower estimates of the HC5 and HC50 values based on the chronic NEC values. No significant differences could be observed between the means and variances of the temperate and Arctic SSDs, resulting in greatly overlapping confidence intervals around the HC5 and HC50 values for temperate and Arctic species. No significant difference in HC5 or HC50

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Table 4 Upper, median and lower estimates of the HC5 and HC50 values derived from the species sensitivity distributions based on NECs (Fig. 2).

Arctic species Temperate species

HC5 median lLower–upper estimate estimate

HC50 medLower–upper ian estimate estimate

0.16 0.05

0.51 0.47

(0.03–0.30) (0.001–0.19)

(0.27–0.97) (0.12–1.86)

values between temperate and Arctic species could be observed. Slope values for the acute SSDs based on the EC50 values are 0.43 and 0.32 for Arctic and temperate species, respectively. Slope values for the chronic SSDs based on the DEBtox estimated chronic NEC values are 0.29 and 0.51 for Arctic and temperate species, respectively.

4. Discussion In the current study we generated new toxicity data on artificial PW in order to elucidate whether temperate data used in risk assessment are sufficiently representative for the Arctic. The selected toxicity parameters EC50 (concentration causing effects to 50% of exposed individuals), NEC (no-effect concentration), HC5 and HC50 values (the hazardous concentrations affecting 5% and 50% of the species) are commonly applied in risk assessment procedures. HC5 and HC50 were calculated from species sensitivity distribution curves developed for Arctic and temperate species for artificial PW. Within the limits of individual species toxicity tests, our data revealed no differences in individual sensitivities between Arctic and temperate species to artificial PW. 4.1. Produced water toxicity Published toxicity data for PW are rare (Lee et al., 2012) and until now, no EC50 values have been reported for Arctic species. The data obtained in the present study provides a valuable addition to the knowledge on tolerance of Arctic species to PW, but given the varying composition of PW, it is difficult to generalize on PW toxicity. Reported EC50 values from tests with PW from various offshore installations around the world, indicate values from 4.66% to 58.8% PW for different taxonomic groups (Holdway, 2002). In the present study, EC50 values for the artificial PW ranged from 18% to 30% for chronic exposure and 36% to 525% for acute exposure. This indicates that our artificial PW showed relative low toxicity, which could possibly be explained by the fact that our PW did not contain production chemicals, which often are added in the production process facilities. Such production chemicals might considerably contribute to the toxicity of PW (Smit et al., 2011). Hence, other results than observed herein can be found when studying effects of produced water in situ. 4.2. Species sensitivities When we compare the order of species sensitivities that was obtained from the tests in the present study there is no clear pattern of specific taxonomic groups being more or less sensitive than others. Some of the Arctic taxa were more sensitive than the temperate ones, and the opposite was observed as well. In addition, the species order based on sensitivity is different when based on EC50 or NEC. Similar results were obtained by Olsen et al. (2011) and De Hoop et al. (2011) who also could not identify particular sensitive taxa. The latter study, however, indicated that Arctic fish species were significantly more sensitive than temperate fish species when exposed to naphthalene. In the present study, this

pattern was confirmed with the two Arctic fishes (B.saida and L. maculatus) species having comparable sensitivities, while the temperate fish (S. maximus) was more robust. The EC50 value of the temperate fish was almost a factor of three higher than for the two Arctic fishes, indicating that the Arctic fish species were more sensitive compared to the temperate fish. In our study the temperate copepod T. battagliai was less sensitive than the temperate algae S. costatum. The same result was obtained in an experiment where T. battagliai and S. costatum were exposed to fluoranthrene for 72 h (Jiang et al., 2010). On the contrary, Brooks et al., (2007) found an opposite relation between EC50 values for the copepod T. battagliai and the algae S. costatum. The difference in toxicity of the artificial PW to the temperate and Arctic species from the same taxonomic group may be due to intrinsic differences such as lipid content, exposure history, and robustness to harsh environmental conditions. For instance, S. costatum can be considered a very tolerant species since it is widely distributed from temperate to certain high Arctic locations (Berge et al., 2005). Furthermore, pairwise comparison may be difficult when temperate and Arctic species of the same taxa have different ecologies. For instance, T. battagliai is a benthic harpacticoid copepod while C. glacialis is a pelagic calanoid copepod with very different life cycle strategies. Also, C. crangon is a shallow water species found on muddy sand, not deeper than 20 m while S. boreas occurs from the littoral zone down to 200–450 m depth on different substrates such as sand or gravel. This species seems to be quite tolerant to environmental factors as it copes seasonally with high loads of finer sediment that may settle during the spring and summer melting season of snow and ice. Finally, S. maximus is a flatfish that lives on sandy, rocky or mixed bottoms, and is commonly found in brackish waters, feeding mainly on bottomliving fish. By contrast, B. saida is a pelagic, or sometimes considered a demersal fish, but feeding on pelagic prey. Hence, direct comparison of toxicity data between temperate and Arctic species of the same taxa may be irrelevant as the selected Arctic species were chosen due to their key role in the Arctic ecosystem and not as comparable homologues in terms of habitat, ecology and biology. This may also explain the lack of trend in sensitivity/robustness of Arctic versus temperate species. Copepods are often indicated as sensitive species but in the present study the Arctic copepod had the highest EC50 value. At the same time, the Arctic copepod also had the lowest NEC of the Arctic species. Consequently the intra species variation in sensitivity was particularly high for this species. In concentrations of 2.27 and 5x PW, a strong narcotic effect was observed on individuals, which made it difficult to determine if an animal was dead or alive. Microscopic observations of each individual were necessary to provide a reliable assessment. The species was sampled in the productive period, with no sea ice and a high abundance of food. In this period, the copepods store energy in big internal lipid droplets, which may also act as sinks for lipid soluble contaminants, where they are stored for the long winter diapause. When stored, the organism may not need to activate energy demanding degradation processes to neutralize the contaminants. Between individual-variation in these processes might have caused higher intra species variation. Among all tested species (both Arctic and temperate), the Arctic algae P. glacialis was the most sensitive species to artificial PW (EC50 ¼ 36%). Hence, for monitoring purposes of potential PW discharges in cold waters, we propose that algae may serve as a good indicator species. Algae are an ecologically important group in most aquatic ecosystems and have previously been shown to be valuable indicators of ecosystem conditions, by responding quickly in both species composition and density to a wide range of water conditions due to changes in water chemistry (e.g. McCormick and Cairns, 1994). Hence, alterations in algae species may

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also provide useful early warning signal for deteriorating conditions in cold water. 4.3. Sensitivity of Arctic vs temperate species assemblages Since pairwise comparison of the species belonging to the same taxa did not reveal any clear patterns, toxicity data have been compiled into SSDs. Species sensitivity distributions are used to set quality criteria and are often believed to represent the sensitivity of species assemblages (Posthuma et al., 2002). The SSDs for the Arctic and non-Arctic species, based on the respective EC50 and NEC values were found to overlap (Figs. 1 and 2). Also the confidence intervals calculated for the HC5 and HC50 values are overlapping indicating that based on single species toxicity data no differences can be observed in species assemblages for Arctic and non-Arctic species. The HC5 based on chronic NECs can be used as a basis for defining a threshold. Based on the available data there is no reason to justify more or less stringent thresholds for Arctic and non-Arctic Environments. However, it should be stressed that this does not mean that environmental impacts are equal in Arctic and non-Arctic environments once thresholds are exceeded. For this purpose responses on higher levels of biological organization need to be studied. The slope values are the standard deviations of the log-normal distributions. A low value equals a steep slope, a high value equals a shallow slope (Smit et al., 2011). The variation in species sensitivity, represented by the slopes of the SSDs, ranges from 0.43 to 0.29 for Arctic species and from 0.32 to 0.51 for the temperate ones. These values do not indicate a higher or lower variation in species sensitivities for the two regions. Harbers et al. (2006) presented a slope value for non-polar narcosis, which is the main toxic mode of action for oil constituents, of 0.65. The value obtained in the present study is lower, which might be caused by the relative limited amount of species that were included. 4.4. Uncertainties in the results Although the present study was carried out with the specific aim to keep inter test variability between the Arctic and the nonArctic exposure tests at a minimum (e.g. harmonized test protocol, dedicated team of experimentalists, etc.) still several sources of variation are included that cause uncertainty in the end result. The section below discusses the most important ones. 4.4.1. Insolubility of chemicals Test protocols were set up to achieve comparable exposure concentrations, following the OECD guidelines in both experiments. However, the insolubility of the chemicals in cold water represented a challenge. For most of the tests, the large volumes of solutions to handle (up to 150 L per treatment) represented a logistical issue that did not allow for using sonication or heating for improving the solubility of the compounds. Furthermore, these large volumes were used to respect the density of organisms in exposure chambers (e.g. 1 g fish per 1 L) as stated in the SOP guidelines. Also, the use of triplicates of the treatments, generally presented an issue for space and handling. In that context, smaller organisms requiring maximum 20 L tanks could be preferred for routine testing. L. maculatus, due to its small size and because it is easy to sample, may be considered as an alternative Arctic test organism instead of B. saida. The measured water concentrations clearly indicated higher solubility in the temperate tests than in the Arctic tests. How the insolubility of chemicals may potentially have affected the toxicity is difficult to assess since the artificial produced water is a complex mixture with a number of different chemicals (Table 1). As we did not study the kinetics of the chemicals in this study, we have not

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assessed how insolubility may have affected the toxicity. As, in the current study, SSD curves were constructed based on measured concentrations for both Arctic and temperate species, we assume that the influence is minimal. However, there should be a discussion to potentially adapt the test protocols for Arctic tests to address solubility issues. 4.4.2. Field collected species versus laboratory cultures Toxicity data may have been influenced by sampling stress. For instance, four Arctic test species (C. glacialis, S. boreas, B. saida and L. maculates) and one temperate species (C. crangon) were collected in the field while the other species came from cultures or farms. Field sampled species may be in a more stressed physiological state, although acclimated to laboratory conditions, and are a more heterogeneous population, including “weak” individuals or different age classes or sex. It has been shown that different life stages of an organism have varying sensitivities to toxicants; early life stages are typically more sensitive than fully developed organisms (Verriopoulos and Moraïtou-Apostolopoulou, 1982; Green et al., 1996). A reduced condition of field collected individuals might be an explanation for high mortality observed in the controls of the test with L. maculatus 27 h after the experiment started and of the unexpected high sensitivity observed for C. crangon. Although mortality of C. crangon in the stock population was less than 10% and the animals did not show any external signs of reduced condition. Overestimated sensitivity of four Arctic species might have influenced the results presented in this paper. The fact that some Arctic organisms such as C. glacialis and M. edulis embryos are not available all year round hampers the routinely application of these assays. In the perspective of implementing routine WETs for Arctic species, protocols to culture the selected polar organisms should be implemented. 4.4.3. Adaptation of test protocols The M. edulis embryo assay was successful under Arctic conditions, although the length of development was significantly longer than for oyster embryos (from 24 h to 300 h). This may have affected the concentration of the chemicals in the water as well as other factors such as oxygenation. However, the effect of these factors on the observed toxicity has not been determined. Considering the long development time from fertilization to Dshaped larva for the M. edulis at environmental temperature (approximately 5 °C), an alternative approach could be the exposure of fertilized eggs and the development of the eggs to the trocophore larvae. The bioassay endpoints could be movement and shape. This alternative should however be tested to determine whether good dose–response relationships would be observed. 4.4.4. Temperature Prosser (1991) proposed that thermal stress acts on the neural membranes, often accompanied by tissue hypoxia. As exposure to toxicants may increase the metabolic oxygen demand, higher temperatures will potentiate the toxic effect. However, since Arctic species were tested under Arctic conditions and non-Arctic species under temperate conditions, thermal stress is not likely to influence the test results. In the present study, the effect of temperature on the biological activity was not evident for most bioassays as toxic effects appeared as rapidly for Arctic species as for temperate species. The slower embryo development rate in the embryo toxicity assay performed with M. edulis offspring was possibly the only temperature effect, increasing the assay time from 24 h to 300 h (discussed above). Indeed, a faster development was observed for embryo samples pipetted from the stock suspension for microscopy observation and left at room temperature overnight. The low EC50 for the acute C. glacialis test may also be partially due to

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low temperature and longer response times and therefore the NEC calculated by DEBtox is a better representative for sensitivity. The effect of temperature (Arctic: 3–9 °C, temperate: 14–24 °C) on the bioassays was primarily evident in the solubility of the chemicals in the Arctic bioassays. The measured water concentrations clearly indicated higher solubility in the temperate tests than in the Arctic tests. Hansen et al. (2011) investigated the effects of biodegraded crude oil on two closely related species of copepods, the temperate C. finnmarchicus kept at 10 °C and the high Arctic C. glacialis at 2 °C. C. glacialis was somewhat less sensitive to oil than C. finnmarchicus, but the difference was small. Toxic effects and mortality occurred a little later in the high Arctic species, and animals with the highest lipid content survived longest. Studies on sensitivity of naupli stages of the same two species to one PAH in combination with increasing temperature, revealed a similar pattern where nauplii of both species were affected by exposure but C. finmarchicus was more sensitive compared to C. glacialis (Grenvald et al., 2013). In the same study, increasing temperature led to increased mortality of C. finmarchicus. Bechmann et al. (2010) also found lower mortality and slower growth in prawn larvae (the cold water species Pandalus borealis) when they were exposed to oil from the Barents Sea at 5 °C than when exposed to North Sea oil at 10 °C. These studies indicate that sensitivity to oil may be somewhat less in Arctic organisms than in marine organisms further south. However, similar experiments conducted on cod indicated no significant difference in toxicity (Jonsson and Björkblom, 2011).

Arctic toxicity data for the derivation of Arctic thresholds seems to be possible, but the assessment of potential impacts in the Arctic based on non-Arctic data should be performed with caution until potential differences in the sensitivity of Arctic and non-Arctic systems at higher levels of biological organization are further studied.

5. Conclusions In this study we investigated the effects of artificial PW on low to high trophic levels from Arctic and temperate habitats. The acute toxicity of artificial PW on Arctic species was similar to that found for the temperate species. We conclude that the Arctic marine species are not less sensitive than their temperate counterparts to artificial PW further confirming the findings of Olsen et al. (2011) when using single species studies. Hence, risk tolerance levels can be derived based on data from both temperate and Arctic species in current risk assessment tools. However, data applied to ecological risk assessment come from single-species toxicity tests measuring effects to individuals only. Generally, it is populations, communities, and ecosystems that are the entities to be protected. To address biological complexity in ecological risk assessment, we recommend improving the knowledge on higher level responses for Arctic populations and communities in order to assess actual impact levels once tolerance levels have been exceeded.

4.5. Implications for risk assessment

Acknowledgement

In order to check the applicability of the existing risk tools to Arctic areas, we have developed basic knowledge on sensitivity of Arctic species to artificial PW, which can be used directly to assess Arctic versus non-Arctic individual sensitivities. Based on the results from the current study and supported by the results from De Hoop et al. (2011) and Olsen et al. (2011) it can be concluded that EC50 tests do not reveal a difference in toxicity on the individual level between Arctic and non-Arctic species. Previously, it has also been shown that the differences are small between sensitivities of polar and temperate species and it can go in both directions. Ferguson (2012) concluded that it should not be expected that Arctic organisms themselves are more sensitive to discharges from the petroleum industry than organisms from other parts of the continental shelf. However, a number of biological and physiological differences between Arctic and temperate systems do exist (e.g. degradation of oil, physical behaviour of oil related to temperature differences, seasonality of feeding and reproductive activities, differences in food availability (AMAP, 1998). Obviously, these factors are not addressed in standard toxicity testing in the laboratory. Previous studies suggest that Arctic communities and populations are more susceptible to chemical exposure (including oil related chemicals) compared to temperate communities (Chapman and Riddle, 2003, 2005; Olsen et al., 2007). And although toxicity thresholds based on individual species toxicity data are likely to be comparable for Arctic and non-Arctic species, actual environmental impacts might still be different once these thresholds are exceeded. Hence, it is important to bear in mind that although on the level of individual species we did not observe differences in this laboratory study, different results than indicated herein may be observed in the field. Until higher level and chronic toxicity studies have been performed including evaluation of developmental stages rather than exposure periods the difference in the response of Arctic and non-Arctic ecosystems will be unpredictable (Hjermann et al., 2007). Hence, the use of non-

This project has been financed by Statoil ASA AS. The writing of the Article was financed by Statoil ASA, COWI, NIVA and Akvaplanniva. We thank the crew of RV Helmer Hansen for sampling organisms in the Arctic and Jørgen Berge for facilitating the work in the laboratory at UNIS, and also the University of Copenhagen for making facilities at the Arctic station, Qeqertarsuaq available. Raw data (raw concentration-response data for all replicates, treatments and species) can be made available upon request.

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