Environmental Toxicology and Chemistry, Vol. 9999, No. 9999, pp. 1–12, 2015 # 2015 SETAC Printed in the USA

Environmental Chemistry EFFECTS OF BENTHOS, TEMPERATURE, AND DOSE ON THE FATE OF HEXABROMOCYCLODODECANE IN EXPERIMENTAL COASTAL ECOSYSTEMS VON STEDINGK,x and KERSTIN GUSTAFSSONz yDepartment of Ecology, Environment and Plant Sciences, Stockholm University, Stockholm, Sweden zSchool of Natural Sciences, Technology and Environmental Studies, S€ odert€ orn University, Huddinge, Sweden xEnvironmental Chemistry Unit, Department of Environmental Science and Analytical Chemistry, Stockholm University, Stockholm, Sweden

CLARE BRADSHAW,yz ANNA STRID,x HANS

(Submitted 6 August 2014; Returned for Revision 1 October 2014; Accepted 16 February 2015) Abstract: The authors studied the fate of the brominated flame retardant hexabromocyclododecane (HBCDD) added in a particulate

suspension to experimental ecosystems assembled from brackish (Baltic Sea) coastal bays. Two experiments examined how benthic macrofauna (over 21 d) and increased temperature (14 d) affected HBCDD concentrations and fractionation of a, b, and g diastereomers in the water, sediment, and biota. A third experiment run over 3 seasons (231 d), studied the effect of HBCDD dose on the same endpoints. In all treatments of the 3 experiments, HBCDD partitioned mainly to the sediment, and this proportion increased with time. Presence of macrofauna tended to increase the HBCDD concentration in the sediment and decreased its concentration in the water. Increased temperature (þ 58C) decreased the amount of HBCDD in sediment and water but not in the filter- and deposit-feeding infaunal bivalves (Macoma balthica). The partitioning between water, sediment, and biota was not concentration dependent. In all treatments, sediment became enriched in g-HBCDD, M. balthica in a-HBCDD, and water in a- and b-HBCDD. Bioaccumulation of HBCDD in M. balthica was high in all experiments (log biota-sediment accumulation factor [BSAF] > 1.25), the a diastereomer contributing the most (log BSAF 2.1–5.2). There is a risk of trophic transfer of HBCDD from benthic to pelagic food webs, as well as secondary poisoning of marine consumers. Environ Toxicol Chem 2015;9999:1–12. # 2015 SETAC Keywords: Environmental fate

Flame retardants

Benthic ecology

Model ecosystems

Bioaccumulation

A main source of HBCDD to the environment is from waste; for example, several recent studies in China have detected high levels of HBCDD around e-waste recycling facilities and dumping sites [5–7]. Textile factories also have been identified as sources [8–10]. Despite the main production of HBCDD being in China, Japan, Europe, and the United States, and its main areas of use being China and Europe, HBCDD has been detected in biota in remote areas such as the Arctic [11–14] and the Tibetan Plateau [15], indicating that long-range transport occurs. Hexabromocyclododecane has a molecular weight of 641.7 g mol1 and log KOW of 5.62. Sixteen possible stereoisomers have been identified; of these, the a, b, and g diastereomers are the most common in the commercially available compound (technical HBCDD), flame-retarded products, and the environment [16]. These diastereomers have different water solubilities (a 48.8 mg L1; b 14.7 mg L1; g 2.1 mg L1), and log KOWs (a 5.07; b 5.12; g 5.47) [3]. Hexabromocyclododecane’s chemical properties mean that in the aquatic environment it partitions mainly to particulate matter and sediment and associates with lipids and biomagnifies in aquatic food webs [6,17–19]. The highest concentrations of HBCDD in wildlife appear to be in species with high trophic levels such as peregrine falcons, guillemots, seals, polar bears, and cetaceans [8,11,14,20,21], thus indicating that biomagnification occurs in nature. Diastereomers of HBCDD fractionate differently in the environment. Sediments tend to be dominated by g-HBCDD, or may be slightly enriched in a [22,23] whereas a-HBCDD dominates in biota [6,17,23–25]. The Baltic Sea is a semi-enclosed, temperate water body stretching from northern Sweden and Finland in the North (65.58N) to Germany and Poland in the South (548N). It is characterized by its shallow depth (average, 56 m), permanent north–south brackish salinity gradient (3–15), cold climate with

INTRODUCTION

Hexabromocyclododecane (HBCDD) is an additive brominated flame retardant used mainly in expanded and extruded polystyrene foam for insulation and construction, as well as in textiles and electric and electronic appliances (high-impact polystyrene). In 2009, the United Nations body, the United Nations Economic Commission for Europe (UNECE) judged HBCDD to fulfill the criteria of persistence, bioaccumulation, and toxicity as part of a screening-level assessment [1], and it is listed as a persistent, bioaccumulating, and toxic substance in the European REACH legislation. Since July 2012, HBCDD has been listed in Annex VI of the European Chemicals Agency (ECHA) Classification, Labeling, and Packaging Regulation as being very toxic to aquatic life under both acute and chronic exposures [2], and in May 2013 it was added to Annex A to the Stockholm Convention, because of its long-range environmental transport, and likely significant adverse human health and environmental effects [3]. It is thus regulated by the Stockholm Convention as a persistent organic pollutant (POP), meaning that all relevant organizations must take measures to globally eliminate its production and use in insulation materials, electronics, and textiles. Building insulation has a time-limited exemption. This, together with apparent continued unregulated use of HBCDD in a range of consumer products [4], means that releases to the environment will continue from these materials. Despite this, knowledge is lacking on how varying environmental conditions affect the fate of HBCDD.

All Supplemental Data may be found in the online version of this article. * Address correspondence to [email protected] Published online 20 February 2015 in Wiley Online Library (wileyonlinelibrary.com). DOI: 10.1002/etc.2947 1

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frequent ice cover during the winter, long water residence time (25–35 y), and low species diversity. The large and heavily populated catchment area (85 million people) means that there is a strong anthropogenic influence, for example, nutrient and chemical inputs. In the Baltic Sea, HBCDD concentrations over the past 3 decades increased steadily in guillemot eggs until a maximum of approximately 200 ng g lipid1 and in cod liver (maximum of 50 ng g lipid1) in 2008, whereas concentrations in herring muscle have been quite constant since the late 1990s, with concentrations ranging between 5 ng g lipid–1 and 50 ng g lipid1, depending on the season and area. Blue mussels have been monitored at 3 sites since the year 2000, and HBCDD concentrations range between 2 ng g lipid1 and 20 ng g lipid1 [26]. However, data on concentrations in other organisms in the Baltic Sea are lacking. Whether the Baltic Sea’s unique properties affect the fate or accumulation in biota of HBCDD also is unknown. Despite HBCDD’s classification as persistent, bioaccumulating, and toxic (REACH) and as a POP (Stockholm Convention), data are severely lacking on partitioning and persistence of HBCDD under realistic environmental conditions, as well as on how the toxic properties of the substance may affect the structure and ecological function of aquatic ecosystems. Therefore, we performed a series of outdoor multispecies experiments in brackish coastal water model ecosystems, aimed at investigating both the ecosystem effects and the fate and distribution of HBCDD in different compartments (water, sediment, benthic organisms) under altered environmental conditions. The results describing the effects on ecosystem structure and function are presented in separate publications. Experiments in model ecosystems are useful as links between less complex laboratory studies that are well controlled but lack environmental relevance, and monitoring data obtained from the field that are environmentally relevant but affected by a natural variability and confounding factors. Although not as complex as natural ecosystems, they add ecological and environmental realism to experimental setups that enable a deepened understanding of the processes that influence the fate and effects of contaminants in the environment, yet also enabling a degree of replicability and control over potentially confounding factors. The present study presents data on the partitioning of HBCDD in 3 experiments that aimed to recreate realistic Baltic Sea shallow coastal ecosystems, with different scales and complexity, and to use an environmentally realistic exposure scenario (HBCDD bound to organic particles in the water column). Experiment A tested whether bioturbating macrofauna influenced the distribution of the test substance between the sediment and water phases. Bioturbation is known to affect benthic-pelagic coupling and the fate of organic contaminants because of mixing, burial, or consumption of contaminated sediments [27,28], but it is often overlooked when assessing partitioning between sediments and water. In experiment B, we investigated how increased temperature (þ58C) affected the distribution of the test substance between the sediment, water, and bivalves. Global warming is expected to increase Baltic Sea water temperatures by an average of 38C to 58C over the next century, with the largest increases in southern areas during the summer [29], and temperature is known to affect both chemical and biological processes. However, the effect of temperature on contaminant fate in natural systems is poorly understood. In experiment C, we investigated whether the fate and distribution of the test substance over 3 seasons, including a period of freezing, was dependent on the initial HBCDD dose applied to the systems.

C. Bradshaw et al. METHODS

General description of the experimental setup (all 3 experiments)

For all experiments, all material (sediment, water, and organisms) was collected from coastal bays in the immediate vicinity of Stockholm University’s Ask€o Laboratory (58849.4’N 17838.2’E), located on the island of Ask€o in the Northern Baltic Proper, 80 km south of Stockholm, Sweden. This area has a low level of human impact and is often used as a reference area in field studies. Thus, levels of HBCDD were expected to be low. All experiments were carried out outdoors with natural light but sheltered from rain. Organism densities used in the experiments were in the range of naturally occurring field densities, with the exception of phytoplankton and zooplankton, which were enhanced. Great care was taken to construct the replicate systems to minimize variability between them. For example, sediment was sieved on a 1-mm mesh to remove macrofauna that were later replaced in equal numbers and similar biomass in all replicate containers. The same volume of water and wet weight of sediment were added from a central well-mixed stock, and the number and wet biomass of each organism type was kept as equal as possible. Nutrient concentrations (dissolved N, P, and Si) in the water were monitored throughout the experiments, and extra Si was added after 35 d in experiment C to avoid depletion in the systems. Details of the 3 setups are given in Supplemental Data, Table S1. Treatments were always randomly assigned to the replicate experimental units. Technical-grade HBCDD mixture (>99.5% purity) was obtained from the former Dead Sea Bromine Group, now known as ICL. In all 3 experiments, HBCDD was added to the experimental systems in a phytoplankton suspension, to mimic a contaminated algal bloom settling out through the water column. Phytoplankton was collected using a plankton net (mesh size 15 mm for experiment A, 90 mm for experiments B and C) in nearby coastal waters. The mesh size was different between the 3 experiments, to optimize collection of the most abundant phytoplankton present at the time of collection. In each experiment, great care was taken to add the same amount (mg organic C) of the suspension to all replicates regardless of treatment and replicate. The amount of suspension added was also chosen to be representative of field conditions. A summary of the nominal amounts of HBCDD in the plankton suspensions are given in Table 1, and an outline of each experiment is given below in Experiment A—The effect of benthic macrofauna on fate of HBCDD in model ecosystems, Experiment B—The effect of temperature on fate of HBCDD in model ecosystems, and Experiment C—The role of exposure on the partitioning of HBCDD in model ecosystems over 3 seasons. Experience from previous brackish water microcosm studies using a similar setup to experiments A and B had shown that the optimal study time was approximately 2 wk to 3 wk, after which the model ecosystems are prone to become unstable (e.g., filamentous algae start to grow in the containers). For experiment C, the mesocosm experiment was planned to run over 3 seasons— autumn, winter, and spring—and lasted 231 d. In the following descriptions, all time points (t) are in days. Experiment A—The effect of benthic macrofauna on fate of HBCDD in model ecosystems

Experiment A investigated the effect of benthic macrofauna on the partitioning of HBCDD between the different

How benthos, temperature, and dose affect HBCDD fate

Environ Toxicol Chem 9999, 2015

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Table 1. Summary of nominal and measured amounts of hexabromocyclododecane (HBCDD) and plankton suspension added to the different treatments

a

Plankton per mL suspension (mg dry wt) % TOC in plankton suspensiona Nominal amount of plankton suspension added to each micro/mesocosma Suspension per L water in micro/mesocosm (mg dry wt) Total organic carbon suspension per L water in micro/mesocosm (mg) Measured HBCDD concentration in plankton suspension added to HBCDD treatmentsb Supernatant (mg g wet wt1) Pellet (mg g dry wt1) Nominal amount of HBCDD added to each micro/mesocosm (mg)b

Experiment A

Experiment B

Experiment C

10.75 38

14 34

10 79–81b

10.0 3.8

16.8 5.7

5.88 4.7

0.73 1610

0.72 3210

5.52, 1.18, 0.49c 35 400, 6560, 956c

0.043

0.079

340, 170, 85, 42.5, 21.3, 10.7, 5.3, 2.7, 1.3d

a

Same amount was added to the control treatments. Values are for spiked treatments only. Data on unspiked suspensions and other measurements in controls are presented in Supplemental Data, Table S2. c The algal suspension was separated into 4 different stock suspensions, of which 3 were spiked with HBCDD (presented here): 1 was added to the 3 highest dose treatments; 1 to the 3 medium dose treatments; 1 to the 3 lowest dose treatments. d The experiment included 9 different HBCDD concentrations and 3 controls. b

environmental compartments of the system, in particular sediment and water phases. The experimental units comprised factory-new 5-L polypropylene buckets (microcosms) approved for food use with approximately 2 cm sediment (sieved on 1 mm) and 4.3 L overlying sand-filtered Baltic Sea water (salinity, 6.5). Phytoplankton used for spiking with HBCDD was collected 29 d before the experimental start (3–4 May 2006), and the resulting 10 L of thick plankton suspension was frozen until spiking several weeks later. The phytoplankton comprised typical spring bloom species (mainly diatoms). Five days before the start of the experiment, the suspension was thawed, thoroughly mixed, and half of the algae suspension was spiked with HBCDD dissolved in an acetone carrier (Table 1), and the other half received the same amount of acetone but no HBCDD. Both suspensions were left at 48C to 58C on a magnetic stirrer until the experiment started. Samples of the suspensions were taken for analysis of HBCDD and the a, b, and g diastereomers. The experiment included 4 treatments in a crossed factorial design: 1) treatment a, with unspiked phytoplankton suspension added, no macrofauna present; 2) treatment b, with HBCDDspiked phytoplankton suspension added, no macrofauna present; 3) treatment c, with unspiked phytoplankton suspension and benthic macrofauna added; and, 4) treatment d, with HBCDD-spiked phytoplankton suspension and benthic macrofauna added. In treatment a and treatment b, all infauna larger than 1 mm had been removed by sieving the sediment. In treatment c and treatment d, the sediment was sieved in the same way, but the following benthic macrofauna were replaced in each replicate: 8 Macoma balthica, 3 Cerastoderma glaucum (both bivalves), and 10 Hydrobiidae (gastropods). The treatments were distributed randomly between 3 troughs through which ambient cool water was pumped to maintain the temperature in the microcosms at 6.5  0.58C, similar to the adjacent natural littoral ecosystem at the time. At the start of the experiment (t0), 4 mL HBCDD-spiked algae suspension was added to each replicate in treatments b and d to achieve a nominal amount of 43 mg HBCDD per microcosm (43 mg dry wt algae or 16.34 mg total organic carbon [TOC]) in each mesocosm (Table 1). The same amount of unspiked algal suspension was added to treatments a and c. The experiment ran for 21 d. Three to 6 replicates of each of the treatments were used. However, only some of these replicates were used for

HBCDD analyses (see Supplemental Data, Tables S4 and S5). The remaining replicates were used for sampling and analysis of effects of the treatments on ecological factors (structure and function of the model ecosystems). Effect data will be presented in a separate publication. Samples for HBCDD analysis of sediment and water were taken from the spiked microcosms (treatments b and d) at the start (t1) and end of experiment (t21), and of M. balthica at t0, t7, and t21, by destructive sampling of certain replicates. The percentage contribution of the 3 main diastereomers (a, b, and g) was also determined in filtered and unfiltered water (t1, t21), sediment (t21), and M. balthica (t7, t21). Well-mixed water samples were siphoned off through a 160-mm mesh to remove zooplankton. Filtration was done later, just before chemical analyses, using grade 3 Munktell filter paper. Sediment samples were taken either from the surface (top 5 mm) or using a minicorer to sample the entire sediment depth (bulk sediment). The M. balthica were washed and allowed to purge their guts for at least 1 h. All samples were frozen immediately at –208C and maintained at this temperature until analysis. Experiment B—The effect of temperature on fate of HBCDD in model ecosystems

Experiment B investigated the effects of increased water temperature on the partitioning of HBCDD between the different environmental compartments, and biota. The experimental units comprised 5-L buckets, of the same type as in experiment A, with 3.75 L sand-filtered Baltic Sea water (salinity, 7.5), approximately 2 cm sediment sieved on a 1-mm sieve and with 4 M. balthica, 2 C. glaucum, and 10 Hydrobiidae added. Phytoplankton used for spiking with HBCDD was collected 7 wk to 8 wk before the experimental start (mid-April 2007), concentrated, and frozen for later spiking with HBCDD. One week before the experiment start, spiked and unspiked suspensions were prepared in the same way as for experiment A. Table 1 summarizes the details of the HBCDD concentrations and amounts of algal suspension used. The treatments were randomly distributed in 2 troughs that maintained the systems at ambient or warm (þ58C) water temperature. The warm treatments were heated using two 300 W aquaria heaters per trough. There were 5 to 6 replicates of each of 4 treatments in a crossed factorial design (ambient

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temperature control, þ58C control, ambient temperature with HBCDD, þ58C with HBCDD). As for experiment A, not all of these replicates were used for HBCDD analyses (see Supplemental Data, Tables S4 and S5). The HBCDD was added as 4.5-mL spiked algal suspension on day 0 to achieve a nominal amount of 78.75 mg HBCDD (63 mg dry wt algae or 16.8 mg TOC) per mesocosm. The same amount of unspiked algal suspension was added to the control buckets. The experiment ran for 13 d. Sediment and water samples for HBCDD analysis were taken from the spiked microcosms at the start (t1) and end of experiment (t13), and of M. balthica at t13, by destructive sampling of certain replicates. The percentage contribution of the 3 main diastereomers (a, b, and g) was also determined in these samples. Samples were taken in the same way as described for experiment A, except that only surface sediment (upper 0.5 cm) was taken.

emptied into a container through a 160-mm mesh. Sediment and benthic bivalve samples were only taken at the end of the experimental period (t231), because sampling during the course of the experiment would have disrupted the sediment. After removing the overlying water and other organisms included in the mesocosms, sediment and M. balthica samples were taken using the same method as described for experiment A. The percentage contribution of the 3 main diastereomers (a, b, and g) was also determined in all samples.

Experiment C—The role of exposure on the partitioning of HBCDD in model ecosystems over 3 seasons

Data handling

The aim of experiment C was to determine whether the partitioning of HBCDD was dose-dependent. In addition, the degree of ecological realism was also increased compared with experiments A and B, by increasing the size and ecological complexity of the experimental setup and running the experiment over 8 mo and 3 seasons (autumn-winter-spring). The mesocosms were 1000-L high-density polyethylene, ultraviolet-resistant, food quality storage containers (Model IBC SLX 1400 from Sotralentz) with the tops partially removed and loose-fitting lids to keep out precipitation while allowing full air circulation. Each mesocosm contained a layer of sediment approximately 8 cm deep and 80 cm (850 L) filtered Baltic seawater (salinity, 6.0). Model ecosystems were constructed by adding organisms representative of typical Northern Baltic littoral ecosystems to each mesocosm, at fieldrelevant densities: 3 types of angiosperms; a macroalga and its associated isopod crustaceans; 2 types of benthic bivalves; and benthic gastropods. Details of numbers, sizes, and densities of these organisms are given in Supplemental Data Table S1. Phytoplankton used for spiking with HBCDD was collected in mid-July 2008, 13 wk before the experimental start, concentrated, and frozen for later spiking with HBCDD. Twelve days before the start of the experiment, spiked and unspiked suspensions were prepared in the same way as for experiments A and B. Table 1 summarizes the details of the HBCDD concentrations and amounts of algal suspension used. In contrast to the 2 previous experiments, a gradient of HBCDD concentrations was used with 3 controls and 9 HBCDD concentrations. HBCDD-spiked phytoplankton suspensions were added to the mesocosms on the 6 October 2008 to achieve nominal amounts of 1.3 mg, 2.7 mg, 5.3 mg, 10.6 mg, 21.3 mg, 42.5 mg, 85 mg, 170 mg, and 340 mg HBCDD per mesocosm. The total amount of phytoplankton added was the same in each mesocosm, including the controls (500 mL, 5 g dry wt, 4 g TOC). Further details are given in Table 1. Treatments were allocated randomly between the 12 mesocosms. The experimental period of 231 d, until June 2009, encompassed the natural fluctuations of environmental conditions induced by the seasonal shifts in the region (autumn, winter, spring), including a period of freezing. Water samples for HBCDD analysis were taken from all 12 mesocosms the day after the HBCDD application (t1), and then at t7 and t231. Samples were taken using a plastic tube that was lowered into the center of the mesocosm to a depth of 40 cm, and

Chemical analyses

Sample preparation, extractions, cleanup, and analyses using gas chromatography mass spectrometry (GC-MS/MS) and liquid chromatography mass spectrometry (LC-MS/MS) were all performed at the Department of Materials and Environmental Chemistry, Stockholm University. Full details of the chemical analyses are given in the Supplemental Data.

Intercomparison of total HBCDD measurements obtained using GC-MS/MS and LC-MS/MS showed that these values were in good agreement (Supplemental Data, Table S1). We thus used them interchangeably to produce the final data set to maximize the amount of data for analyses. Hexabromocyclododecane in the different compartments was compared both as total concentrations and as a percentage of the total amount of HBCDD in each mesocosm. The latter was calculated in 2 ways: either taking into account abiotic compartments only, to allow comparisons between treatments with and without macrofauna; or including benthic bivalves, allowing an estimation of how much of the total amount of HBCDD was accumulated by these organisms (see Supplemental Data). Diastereomer composition was expressed as percentage point change relative to the technical product. Calculation of biota–sediment accumulation factors (BSAF)

Biota–sediment accumulation factor values were calculated for M. balthica, the only species that was sufficiently abundant to provide enough biomass for HBCDD analysis. Biota– sediment accumulation factors were calculated for both total HBCDD and for each of the 3 measured diastereomers. For details, see the Supplemental Data. Statistical analyses

For experiment A, only single HBCDD measurements per treatment were available; thus, no statistical analysis was possible. For experiment B, a 2-way analysis of variance (ANOVA; factors: time, temperature) was performed on HBCDD concentrations in sediment, and a 3-way ANOVA (factors: time, temperature, filtration) on water concentrations. A 2-way ANOVA was performed on the percentage point change in diastereomer composition in sediment (factors: temperature, time) and water at t1 (factors: temperature, water fraction). For experiment C, linear regressions were fitted to the log-transformed concentration data for water (3 time points, unfiltered and filtered), sediment (t231), and M. balthica (t231). The slopes of these regressions were tested for deviation from zero using a t test. To determine whether regression slopes varied depending on time and between filtered and unfiltered water, 2-tailed Z tests were performed for t1 and t7 data. For the 3 diastereomers, linear regressions were performed separately for a, b, and g for water (t1, t7, unfiltered, filtered), sediment, and M. balthica to assess whether a change in fractionation occurred with increasing amounts of HBCDD added. Slope

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Environ Toxicol Chem 9999, 2015

present (Figure 1A, B [t21]). A clear accumulation of HBCDD by M. balthica was seen by t7 (Figure 1C). When considering only the abiotic components of the microcosms, most (>99%) of the HBCDD was found in sediment after 21 d (Table 2). Of the small percentage found in the water, at least twice as much was found in the particulate fraction compared with the dissolved phase. The presence of macrofauna also slightly reduced the percentage of HBCDD in the water and slightly increased the particulate to dissolved ratio. Where infaunal bivalves were present and included in the calculations, these accumulated as much as 10.9% of the total measured HBCDD (macrofauna treatment, t21). Experiment B (effect of increased temperature). Significant effects of time, temperature, and filtration on HBCDD water concentrations were seen (see Supplemental Data for full statistical results). Unfiltered water had a higher HBCDD concentration than filtered, concentrations in both fractions decreased from t1 to t13, and warmer treatments (þ58C) had lower HBCDD concentrations than ambient treatments (Figure 1D). Significant 2-way interactions occurred between all factors—temperature and filtration effects decreased with time, and unfiltered water concentrations were more influenced than the filtered concentrations by temperature (Figure 1D). For sediment, no significant effect of either temperature or time was seen, although a tendency for lower HBCDD concentrations was seen at higher temperature (p ¼ 0.099; Figure 1E). For M.

coefficients (‘a’) were tested for significant deviation from zero using the t test described above. For full details of all statistical tests, see the Supplemental Data. RESULTS

A summary of environmental conditions (temperature, salinity, oxygen levels, pH, nutrients) in the 3 experiments are given in the Supplemental Data, Table S2. Background concentrations of HBCDD in the nonspiked treatments or time points were low or below detection limit (Supplemental Data, Table S3). More than 95% of the HBCDD mixed into the phytoplankton suspensions adsorbed to the particulate phase (“pellet” in Table 1). Fate of HBCDD in the experimental systems

Experiment A (effect of benthic macrofauna). Because of the low number of replicates available for the HBCDD sampling, no statistical tests could be performed, so the results here are purely descriptive. Unfiltered water had higher concentrations of HBCDD than filtered water, and the concentration of both fractions decreased over the 3-wk experiment (Figure 1A). Sediment concentrations showed a large increase from t1 to t21 (Figure 1B). A tendency also was seen for HBCDD concentrations to be lower in water (both dissolved and particulate fractions) and higher in sediment where macrofauna were

A) Water

μg gdry weight

1.E-04 1.E-05

*

1.E-06

1000

0.1

100

μg g lipid-1

filtered

1

-1

unfiltered

1.E-03

μg L-1

Experiment A

C) M. balthica

B) Sediment

1.E-02

0.01

10

0.001

1 t0

E) Sediment

1.E-02

1

unfiltered filtered

e f

1.E-04

g f

1.E-05

t1, ambient

t1, +5C

t13, ambient

t13, +5C

t21

1000

0.1

0.01

1.E-06

t7

F) M. balthica

μg g lipid-1

1.E-03

μg L-1

Experiment B

c d

μg g dry weight -1

D) Water a b

5

#

100

10

1 t13, t13, +5C ambient

Figure 1. Measured concentrations of hexabromocyclododecane (HBCDD) in water, sediment, and the bivalve Macoma balthica in experiments A (effect of benthic macrofauna) and B (effect of increased temperature) at various time points. Where more than 1 sample was analyzed, average values ( standard error) are shown. # ¼ below limit of detection; * ¼ no sample analyzed. Note the different scales on the y-axes and that the y-axes have a log scale. ‘t’ ¼ time in days since addition of HBCDD. In (A) and (B), þM and –M are treatments with and without macrofauna, respectively. Different letters in (D) indicate significant differences between treatments (2-way analysis of variance with Fisher’s least significant difference post hoc test).

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Table 2. Hexabromocyclododecane (HBCDD) partitioning expressed as percentage in different compartments As percentage of total in the system Experiment/time (t, days)/treatment Experiment A t1 no macrofauna t1 with macrofauna t21 no macrofauna t21 with macrofauna Experiment B t1 þ 58C t1 ambient t14 þ 58C t14 ambient Experiment Cb 170 mg 85 mg 42.5 mg 2.7 mg 1.3 mg

a

As percentage of the abiotic components

Bivalves

Sediment

Particulate

Dissolved

Sediment

Particulate

Dissolved

 NA – 10.92

– NA – 88.83

– NA – 0.18

– NA – 0.07

8.27 NA 99.19 99.72

62.0 NA 0.51 0.20

29.73 NA 0.31 0.08

NA NA 13.60 9.36

NA NA 86.21 90.58

NA NA 0.14 0.04

NA NA 0.04 0.02

92.96 92.76 99.79 99.93

5.83 6.05 0.17 0.05

1.22 1.18 0.04 0.02

0.25 0.17 0.15 0.68 3.60

99.16 99.77 99.49 97.91 95.11

0.31 0.04 0.08 0.77 0.89

0.28 0.03 0.28 0.65 0.41

99.41 99.94 99.64 98.57 98.66

0.31 0.04 0.08 0.77 0.92

0.28 0.03 0.28 0.65 0.42

a

First 4 columns are expressed as a percentage of the total amount in the experimental system, last 3 columns as a percentage of the abiotic part (sediment, particulate, dissolved) only. t231; data for only 5 of the treatments available. NA ¼ data not available; – ¼ no bivalves present in this treatment. b

balthica, only 1 value was available for each treatment (t13 ambient, t1 þ 5C), so no statistical comparison was performed. However, no obvious effect of temperature was found (Figure 1F). When considering only the abiotic components of the microcosms, most (92–93%) of the HBCDD was found in the sediment, even after only 1 d. This increased to greater than 99% after 13 d (Table 2). No obvious effect of temperature on the partitioning in the sediment or water was seen. If infaunal bivalves were included in the calculations, these accounted for 9.4% and 13.6% of the total measured HBCDD (t13, þ58C and ambient, respectively). Experiment C (effects of HBCDD dose over 3 seasons). Concentrations of HBCDD in all compartments increased with increasing dose added to the mesocosms (Figure 2, Table 3). The slopes of the regressions were statistically different from zero for filtered and unfiltered water at t1 and t7, sediment, and M. balthica. The slopes were not significantly different from zero for unfiltered and filtered water at t231. A significant difference was found between the slopes of the regressions for filtered and unfiltered water at t1 (Z ¼ 2.74, p < 0.05) and t7 (Z ¼ 2.06, p < 0.05), as well as between unfiltered water at t1 and t7 (Z ¼ 13.78, p < 0.05). No significant difference was found between the slopes of the regressions of filtered water at t1 and t7. After 231 d, more than 98% of the HBCDD was found in the sediment, and the remaining small percentage in the water was divided roughly equally between the particulate and dissolved phase (when considering only the abiotic components of the microcosms; Table 2). No apparent relationship was found between the amount of HBCDD added to mesocosms and the percentage of partitioning after 231 d. When bivalves were included in the calculations, they accounted for only a few percent of the total HBCDD. Diastereomers

The average ( standard error [SE]) percentage contribution of the a, b, and g diastereomers in the technical product that was used to spike the algal suspensions was 2.1  0.2; 8.2  2.6; and 89.7  2.4, respectively. All of the following

results are expressed as percentage point differences from these values. Experiment A (effect of benthic macrofauna). No statistical tests were possible, but some trends can be seen. Filtered water was more depleted in g-HBCDD and more enriched in a and b than unfiltered water, compared with the technical product added to the microcosms (Figure 3A). Sediment containing macrofauna was more depleted in b and more enhanced in g than sediment without macrofauna (Figure 3B). Macoma balthica became depleted in g and enriched in a (and to some degree b) over the 21 d of the experiment (Figure 3C). Experiment B (effect of increased temperature). One day after the addition of HBCDD to the microcosms (t1), significant differences were found in diastereomer composition between the filtered and unfiltered water fractions (Figure 3D), with a and b diastereomers being significantly enriched in the filtered water and g significantly depleted. A significant increase in a and decrease in g also was seen in the water of the warmer (þ58C) treatments. No significant interactions were found between temperature and water fraction. For water at t13, no statistics were possible, nor any obvious trends observed. In the sediment, g-HBCDD was enriched and b depleted, but no significant effect of time or temperature was seen on any of the diastereomers (Figure 3E). Macoma balthica were strongly depleted in g-HBCDD and enriched in a, but this effect seemed slightly less clear in bivalves in warmer water (Figure 3F). Experiment C (effects of HBCDD dose over 3 seasons). The only regression slopes to be significantly different from zero were for t1 filtered water; the a diastereomer had a positive slope (a ¼ 0.016, t-statistic ¼ 2.36, significant at 0.05) and g a negative slope (a ¼– 0.034, t-statistic ¼ 2.36, significant at 0.05) with increasing amounts of HBCDD added (Figure 4A). For the rest of the water data, a change in pattern occurred from t1 to t7 for unfiltered water, from being depleted in b and enriched in g to the reverse (Figure 4B), presumably because of particles settling out (t1 unfiltered has a similar diastereomer profile to sediment because it contains a lot of particles). Sediment was enriched in g and depleted in b, and M. balthica was strongly enriched in a and strongly depleted in g (Figure 4B).

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Table 3. Summary of linear regression coefficients and statistics for hexabromocyclododecane (HBCDD) concentrations in experiment C

A) Water 0

log HBCDD concentration (μg ml-1)

Compartment -1

t1, unfiltered water t1, filtered water t7, unfiltered water t7, filtered water t231, unfiltered water t231, filtered water t231, sediment t231, Macoma balthica

-2

-3

-4

-5

0

0.5

1

1.5

2

2.5

3

B) Sediment log (HBCDD concentration (μg g dry wt -1)

b

R2

0.946 0.642 0.742 0.545 0.610 0.662 1.085 0.927

–3.390 –3.618 –3.893 –3.914 –5.245 –5.686 –1.270 0.145

0.981* 0.800* 0.967* 0.783* 0.690 0.722 0.954* 0.946*

* Slope significantly different from zero (p < 0.05). t ¼ number of days since start of experiment.

-6

2 1.5 1 0.5 0 -0.5 -1

accumulation (log BSAFs 2.1–5.2) compared with the other 2 isomers (b < 1.8, g < 0.4) and with log BSAFs calculated from total HBCDD concentrations (Figure 5). Log BSAFs were higher in experiment C (2.9–3.9) than in the other 2 experiments (1.3–1.5; Figure 5), suggesting a possible increase with time. The same trend was seen for a-log BSAFs (experiment C, 4.5– 5.2; experiments A and B, 2.1–2.4; Figure 5). Both b- and gHBCDD log BSAFs were fairly similar across all experiments. Increased temperature may have increased log BSAF (experiment B: þ58C treatment, 1.5; ambient treatment 1.3) and a-log BSAF (experiment B: þ58C treatment, 2.4; ambient treatment 2.1). However, no significant relationship was found between diastereomer-specific log BSAFs and HBCDD concentration in the sediment (linear regression: slopes not different from zero). DISCUSSION

-1.5 0

0.5

1

1.5

2

2.5

3

C) M. balthica 3

log (HBCDD concentration (μg g lipid-1)

a

2.5

2

1.5

1

0.5

0 0

0.5

1

1.5

2

2.5

3

log (nominal amount HBCDD added to mesocosm (mg))

Figure 2. Experiment C (effect of hexabromocyclododecane [HBCDD] dose over 3 seasons): HBCDD concentrations measured in (A) water, (B) sediment (after 231 d), and (C) the bivalve Macoma balthica (after 231 d), as a function of the nominal amount HBCDD added to the mesocosms. In (A), squares are measurements after 1 d, triangles after 7 d, and circles after 231 d. Solid symbols are unfiltered water samples; empty symbols are filtered water. Details of the regressions are given in Table 3.

BSAF (Experiments A, B, and C)

A strong bioaccumulation (log BSAF > 1.25) of HBCDD occurred in M. balthica in all experiments (Figure 5). Diastereomer-specific BSAFs showed that a has a much higher

Persistence and fate of HBCDD

Information on the persistence of HBCDD in ecosystems is sparse and sometimes contradictory. The European Chemicals Agency (EC) Risk Assessment Report for HBCDD [3] concluded that although some reliable studies indicate that biodegradation can occur, HBCDD does not degrade rapidly. For example, the report presents data on half-lives (DT50; the period required for 50% degradation or dissipation) for aerobic freshwater sediments of 101 d and 191 d (208C and 128C, respectively), and 66 d and 125 d for anaerobic sediments at the same temperatures. These estimations are based on data from controlled laboratory experiments, sometimes including depuration periods in uncontaminated media, with no input of HBCDD after the initial addition, and without biota, so their applicability to field situations may be limited. However, monitoring data indicate a significant degree of environmental transport and stability [3], since HBCDD is found in places far from any point source (e.g., the Arctic [11–14]). The present study did not determine chemical half-life but demonstrated that HBCDD is persistent under realistic environmental conditions. In all experiments, HBCDD concentrations in sediment and M. balthica increased and water concentrations decreased with time, but HBCDD was still easily detectable in all compartments even after 231 d (8 mo; experiment C). Concentrations measured in the 3 compartments were also directly proportional to the amount of HBCDD added to the experimental systems. Experiment C was performed from autumn to spring, and temperatures were generally below 148C, including a period of freezing. Mobilization and degradation processes are thus likely to be slower under these natural conditions. In shallow coastal bays around the Baltic Sea, the focus of these experiments, the water may be frozen, often to the seabed, during the winter.

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C. Bradshaw et al.

B) Sediment

A) Water 10

Experiment A

5

80

4

60

3

5

40

2 1

%

C) M. balthica

0

0

0

-1

-20

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20

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-80 t7

D) Water 35

a a

E) Sediment b b

c ab

d b

25

t21

F) M.balthica

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0

0

-5

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-15

-30

Experiment B

15 5 % -5 -15 -25 -35

a

bc

ab

c

Figure 3. Percentage point change in diastereomer composition, relative to technical mixture added to the experiments. Experiment A (effect of benthic macrofauna): (A) different fractions of water at t1 and t21; (B) bulk sediment at t21; (C) the bivalve Macoma balthica at t7 and t21. Experiment B (effect of increased temperature): (D) different fractions of water at t1 and t14; (E) surface sediment; (F) M. balthica. Where there were multiple replicates, average values ( standard error) are shown. Black ¼ a diastereomer, hatched ¼ b, white ¼ g. Note that y-axes have different scales. Different letters in (D) indicate significant differences between treatments (2-way analysis of variance with Fisher’s least significant difference post hoc test); from top to bottom, results are for a, b, and g.

Shallow water temperatures also frequently reach greater than 258C on a hot summer day. Increasing the temperature of the systems by 58C (experiment B) decreased the total amount of HBCDD detected in all compartments. This could either be attributable to the volatility of the compound increasing, leading to dissipation from the experimental systems to the air, or an increase in degradation rate of HBCDD, through either chemical or microbial processes. However, based on the chemical properties of HBCDD, for example, low vapor pressure and high adsorption potential to suspended matter [3], evaporation from the water surface of the experimental chambers or from the natural coastal zone environment is unlikely to be an important route of dispersion.

All 3 diastereomers persisted throughout the experiments, and fractionation was generally not dose dependent (experiment C). In all experiments, sediment was depleted in b and enriched in g, as found in previous studies [22,23]. Water was generally depleted in g and enriched in a and b. However, these trends were less pronounced in the unfiltered water, presumably because these samples contained particles that would have had a more similar diastereomer profile to sediment. This conclusion is reinforced by the change in fractionation pattern from t1 to t7 in experiment C, which is probably caused by particles settling out from the water to the sediment during this time. Increased temperature initially (t1) increased the fractionation of the a and g diastereomers (experiment B).

How benthos, temperature, and dose affect HBCDD fate

A)

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B)

25

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15 10

-10 -15 -20 -25 1

10

100

Nominal amount HBCDD added to mesocosm (mg)

Figure 4. Experiment C (effect of hexabromocyclododecane [HBCDD] dose over 3 seasons): percentage point change in diastereomer composition, relative to technical mixture added to the experiments. (A) Specific data for unfiltered water at t1, where percentage change in a and g diastereomers changed with dose added to the mesocosm. (B) Water (filtered and unfiltered at t1, t7, and t231), sediment and the bivalve Macoma balthica; averages and standard error of all treatments combined. In (B), filled circles ¼ a diastereomer, crosses ¼ b, open circles ¼ g. Note that the y-axes have different scales. In (A), black ¼ a, hatched ¼ b, white ¼ g.

Most experimental studies apply HBCDD in solution and thus quickly reach the maximum solubility of 65.6 mg L1. However, in the environment, HBCDD is rarely present in the dissolved form, but rather sorbed to suspended particles, soils, and sediment. We therefore chose the experimentally novel and more realistic scenario of adding particle-bound HBCDD, which also allowed us to add as much as 400 mg (in particulate form) L1 (experiment C). The HBCDD application was thus performed simulating a natural route of exposure relevant for limnic, brackish water, and marine ecosystems, that is, where particulate organic matter (settling dead phytoplankton blooms) scavenges the water column for hydrophobic contaminants on its way to the lake or sea floor. Infaunal bivalves and bioaccumulation

logBSAF (μg g lipid-1) / (μg g orgC-1)

Infaunal macrofauna, in particular the filter-feeding and deposit-feeding bivalves that were the dominant species in these 6 5 4 3 2 1 0 -1 -2 0,1

1

10

100

1000

HBCDD concentration in sediment (μg g orgC-1) Figure 5. Log biota-sediment accumulation factors (BSAF) for Macoma balthica at the end of the experiments (experiment A—21 d; experiment B—14 d; experiment C—231 d) plotted against measured sediment concentrations. Small symbols are experiments A and B; large symbols are experiment C. Triangles ¼ total hexabromocyclododecane (HBCDD), filled circles ¼ a diastereomer, crosses ¼ b, open circles ¼ g.

systems, decreased HBCDD concentrations in the water and increased the amount of HBCDD that is sequestered in the sediment (experiment A), presumably by filtering out particlebound HBCDD in the water column and depositing it as feces or pseudofeces in or on the sediment. Benthic macrofauna also enhanced the depletion of b-HBCDD and enrichment of g (experiment A), probably through bivalve bioturbation changing sediment properties such as organic carbon and water content [28]. The importance of bivalves in benthic-pelagic coupling is well known in ecology, and the results in the present study are in line with previous studies that have shown the importance of macrofauna for the fate, recirculation, and bioavailability of hydrophobic organic contaminants [27,28]. The bivalves themselves also took up substantial amounts of HBCDD during these experiments. Our BSAF values (20–30 for the shorter experiments, >1000 for the longer experiment) are the highest of those reported in (or calculated from) the literature: 2.5 to 7.9 for a-HBCDD and 0.3 to 1 for g-HBCDD for freshwater molluscs downstream of a textile factory [10]; 0.6 to 15 for pike in a Swedish river affected by textile industries [30]; less than 1 in flounder in a 78-d experiment [31]; less than 1 for a range of species near an e-waste site [6]; less than 1 for marine invertebrates downstream of a polystyrene factory [23]. Our test species, M. balthica, is both a deposit feeder and a filter feeder and will have eaten newly settled HBCDD-contaminated high-quality food particles on the sediment surface. Species with this feeding behavior are likely to have the highest accumulation of sediment-associated contaminants of all benthic organisms, and contaminants associated with organic matter are very bioavailable to such organisms [32]. The HBCDD concentrations in M. balthica in these experiments were 1 mg g lipid1 to 250 mg g lipid1 (0.05– 10 mg kg wet wt1), with a strong linear correlation between their body concentrations and sediment concentrations. Field concentrations for M. balthica are not available, but values of up to 17 mg g lipid1 have been measured in the filter-feeding blue mussel (Mytilus edulis) [3,33]. Smolarz and Berger [34] exposed M. balthica to HBCDD concentrations in the same range as our higher exposures in experiment C and used the same method of application (in algal suspension). They observed an increase in nuclear abnormalities and in the

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Environ Toxicol Chem 9999, 2015

frequency of dead cells, suggesting toxicity at these doses. This not only is important for the bivalves themselves, but it also has important implications for their vital role in benthic-pelagic coupling and biogeochemical cycling in the sediment. Macoma balthica are a dominant part of soft sediment habitats in the Baltic Sea and also provide an important food source to other benthic invertebrates and bottom-feeding fish [35]. The high accumulation potential of HBCDD by M. balthica thus provides a pathway for HBCDD to be transferred in the food web, potentially causing secondary poisoning to higher trophic levels. The M. balthica was clearly enriched with the a diastereomer and depleted in g, both in terms of percentage of change from the technical product and in terms of diastereomer-specific log BSAFs, in agreement with previous field measurements [6,10,17,23–25]. Although g-HBCDD log BSAFs are comparable between the present study and La Guardia et al. [10], a-HBCDD values in the present study (4.5– 5.2 in experiment C) are considerably higher than those found by La Guardia et al [10] (0.5–1) for the benthic bivalve Corbicula fluminea. These trends in fractionation may be caused by differences in either bioavailability or assimilation efficiency between the diastereomers [36,37], the fact that aHBCDD is more persistent and less easily degraded than b and g, or because organisms bioisomerize b and g forms to a or metabolize them to other molecules [10,17,25,36,38]. We showed that little difference existed in diastereomer fractionation during the first week (experiment A) but that a enrichment in biota seems to increase with time of exposure (20% enriched after 2 to 3 wk in experiments A and B, but 70% after 8 mo in experiment C). In a laboratory experiment with carp, Zhang et al. [38] showed a similar increase in the percentage of a-HBCDD over a period of 50 d. We also demonstrated that fractionation in M. balthica does not seem to be dose-dependent within the range of doses tested (experiment C). This is in contrast to Du et al. [36], who found concentration-dependent bioaccumulation of the 3 diastereomers from contaminated food in a laboratory experiment with zebrafish. Environmental Risk Assessment

The measured concentrations in the surface sediment of the experimental chambers can be compared with available monitoring data from sediments at contaminated sites, and with predicted environmental concentrations used for the environmental risk assessment of HBCDD. In experiments A and B, the measured sediment concentrations in the exposed systems reached 100 ng g dry weight1 to 380 ng g dry weight1. In experiment C, the measured sediment concentrations in the exposed mesocosms varied between 0.1 ng g dry weight1 and 42.7 ng g dry weight1. This means that in experiments A and B and in the 5 lowest treatments of the 9 HBCDD-dosed mesocosms in experiment C, the measured concentrations (100–1180 ng g dry wt1) are in the same range as reported from contaminated sites in Europe (54–1680 ng g dry wt1 [25]) but higher than those reported in sediments from an e-waste dismantling site in China (4.6–35 ng g dry wt1 [7]). Our values are also in the same range as predicted local environmental concentrations (PEClocal) for freshwater and marine sediments that may result from HBCDD production and industrial use [3] and electronic waste treatment [39]. The measured concentrations in the surface sediment and water of the experimental chambers also can be compared with ecotoxicity effect concentrations used for the environmental risk assessment in the ECHA Risk Assessment Report for HBCDD [3]. The highest measured sediment concentrations

C. Bradshaw et al.

(Mexp.sedC) in the present study can be expected to present a risk to aquatic organisms, if compared with the predicted no effect concentration (PNEC) values for freshwater (0.86 mg g dry wt1) and marine (0.17 mg g dry wt1) sediments given in the ECHA Risk Assessment Report for HBCDD [3]. The risk quotient (Mexp.sedC/PNEC) obtained is greater than 1 at the end of experiment B and for all except the lowest 2 exposures in experiment C, when using the marine sediment PNEC value [3]. Using the freshwater PNEC value gives fewer values greater than 1. The risk ratios obtained when comparing the highest measured experimental water concentrations (Mexp.waterC) in our experiments with the PNEC values presented above are never greater than 1, even 1 d after the addition of HBCDD, when using filtered water concentrations (marine and freshwater). However, when using unfiltered water concentrations, which are more relevant for filter feeding organisms, the ratio exceeds 1 in the 2 highest exposures of experiment C (marine PNEC used), indicating cause for concern. No PNEC values specific for brackish environments are presented in the ECHA Risk Assessment Report for HBCDD [3]; therefore uncertainties are associated with assessing the potential environmental risks of this, and many other substances, for Baltic Sea organisms. Applying the ECHA standard procedure for assessment of risk of secondary poisoning [40], we compared our measured concentrations in M. balthica (Mexp.biotaC values) with the PNECoral value of 5.0 mg kg fresh weight1 used in the ECHA Risk Assessment Report for HBCDD [3]. The risk quotient obtained is less than 1 in all cases except the 2 highest treatments in experiment C, where values of 1 and 2 are obtained. However, this approach to assess the risks for secondary poisoning via the aquatic food chain presents a simplistic scenario developed for the schematic food chain water ! aquatic organism ! fish ! fish-eating bird or mammal, using a PNECoral value based on laboratory studies with rats [3]. This simplified food chain is only 1 example of a secondary poisoning pathway, and safe levels for fish-eating animals do not exclude risks for other birds or mammals feeding on other aquatic organisms (e.g., mussels and worms), or, in the case of using exposure data from the present experimental studies, the risks of trophic transfer and secondary poisoning between benthic organisms (M. balthica) and implications for their natural predators (e.g., fish). CONCLUSIONS

The experiments described in the present study have demonstrated the role of bioturbating macrofauna, which accumulated significant amounts of the added HBCDD and also altered the distribution of HBCDD between water and sediment by increasing the amount in the latter. In contrast, increased temperature decreased the amount of HBCDD in the sediment and in the water but not in filter- and deposit-feeding infaunal bivalves. Furthermore, the persistence of the test substance was demonstrated in a long-term study lasting over 3 seasons, in which the test substance was detected in all concentration treatments even 231 d after the initial application. However, neither the partitioning of total HBCDD nor its 3 dominant diastereomers seemed to be concentration dependent. In all treatments of all experiments, the sediment became enriched in g-HBCDD, M. balthica enriched in a-HBCDD, and water in a- and b-HBCDD. The concentrations applied and measured in the various compartments of our 3 experiments were in the same range as

How benthos, temperature, and dose affect HBCDD fate

those measured at contaminated sites, and in the same range as those predicted (predicted environmental concentrations) in the European risk assessment for HBCDD [3] and the risk assessment performed under the Restriction of Hazardous Substances Directive [39]. At these concentrations, the risk of toxicity to aquatic organisms from the sediment and of secondary poisoning to organisms feeding on the benthic bivalve M. balthica cannot be excluded. Environmentally realistic experimental setups allowed us to demonstrate the role of ecology and environmental factors on partitioning of HBCDD and to determine the possibility of ecotoxicological risk in such systems, with implications for risk assessment based on field conditions. SUPPLEMENTAL DATA

Tables S1–S5. (189 KB PDF). Acknowledgment—We gratefully acknowledge excellent assistance with these labor-intensive experiments from A. Cala, E Håkansson, N. Olsson, H. von Seth, and the staff of the Ask€ o Laboratory, especially M. Murphy and S. Ericsson. I. Athanassiadis, Department of Materials and Environmental Chemistry, Stockholm University, is thanked for performing the GC-MS/ MS analysis. The present study was funded by grants from the Baltic Sea Foundation and from King Carl XVI Gustaf’s Foundation for Research and Education and by Stockholm University’s Baltic Ecosystem Adaptive Management (BEAM) program. Data Availability—Most data are available in the Supplemental Data section. Any other information may be obtained from the lead author (clare. [email protected]). REFERENCES 1. Marvin CH, Tomy GT, Armitage JM, Arnot JA, McCarty L, Covaci A, Palace V. 2011. Hexabromocyclododecane: Current understanding of chemistry, environmental fate and toxicology and implications for global management. Environ Sci Technol 45:8613–8623. 2. ECHA (European Chemicals Agency). 2014. Online Classification and Labelling Inventory Database. Helsinki (Finland). [cited 2014 June 7]. Available from: http://echa.europa.eu/web/guest/information-onchemicals/cl-inventory-database 3. ECHA (European Chemicals Agency). 2008. Risk assessment: Hexabromocyclododecane. Final Report May 2008. ECB, Luxembourg. 4. Rani M, Shim W, Han GM, Jang M, Song YK, Hong SH. 2014. Hexabromocyclododecane in polystyrene based consumer products: An evidence of unregulated use. Chemosphere 110:111–119. 5. Eguchi A, Isobe T, Ramu K, Tue NM, Sudaryanto A, Devanathan G, Viet PH, Tana RS, Takahashi S, Subramanian A, Tanabe S. 2013. Soil contamination by brominated flame retardants in open waste dumping sites in Asian developing countries. Chemosphere 90:2365–2371. 6. Wu J-P, Guan Y-T, Zhang Y, Luo X-J, Zhi H, Chen S-J, Mai B-X. 2010. Trophodynamics of hexabromocyclododecanes and several other nonPBDE brominated flame retardants in a freshwater food web. Environ Sci Technol 44:5490–5495. 7. Zhang X, Yang F, Luo C, Wen S, Zhang X, Xu Y. 2009. Bioaccumulative characteristics of hexabromocyclododecanes in freshwater species from an electronic waste recycling area in China. Chemosphere 76:1572–1578. 8. Sellstr€om U, Bignert A, Kierkegaard A, Haeggberg L, de Wit CA, Olsson M, Jansson B. 2003. Temporal trend studies on tetra- and pentabrominated diphenyl ethers and hexabromocyclododecane in guillemot egg from the Baltic. Environ Sci Technol 37:5496–5501. 9. Chen D, La Guardia MJ, Luellen DR, Harvey E, Matteson Mainor T, Hale RC. 2011. Do temporal and geographical patterns of HBCD and PBDE flame retardants in U.S. fish reflect evolving industrial usage? Environ Int 45:8254–8261. 10. La Guardia MJ, Hale RC, Harvey E, Matteson Mainor T, Ciparis S. 2012. In situ accumulation of HBCD, PBDEs, and several alternative flame-retardants in the bivalve (Corbicula fluminea) and gastropod (Elimia proxima). Environ Sci Technol 46:57985805. 11. de Wit CA, Alaee M, Muir DCG. 2006. Levels and trends of brominated flame retardants in the Arctic. Chemosphere 64:209–233.

Environ Toxicol Chem 9999, 2015

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12. de Wit CA, Herzke D, Vorkamp K. 2010. Brominated flame retardants in the Arctic environment: Trends and new candidates. Sci Total Environ 408:2885–2918. 13. Dietz R, Riget FF, Sonne C, Born EW. 2013. Three decades (1983– 2010) of contaminant trends in East Greenland polar bears (Ursus maritimus). Part 2: Brominated flame retardants. Environ Int 59: 494–500. 14. Hoguet J, Keller JM, Reiner JL, Kucklick JR, Bryan CE, Moors AJ, Pugh RS, Becker PR. 2013. Spatial and temporal trends of persistent organic pollutants and mercury in beluga whales (Delphinapterus leucas) from Alaska. Sci Total Environ 449:285–294. 15. Zhu N, Fu J, Gao Y, Ssebugere P, Wang Y, Jiang G. 2013. Hexabromocyclododecane in alpine fish from the Tibetan Plateau, China. Environ Pollut 181:7–13. 16. Heeb NV, Graf H, Schweizer WB, Lienemann P. 2010. Thermallyinduced transformation of hexabromocyclododecanes and isobutoxypenta bromocyclododecanes in flame-proofed polystyrene materials. Chemosphere 80:701708. 17. Law L, Halldorson T, Danell R, Stern G, Gewurtz S, Alaee M, Marvin C, Whittle M, Tomy G. 2006. Bioaccumulation and trophic transfer of some brominated flame retardants in a Lake Winnipeg (Canada) food chain. Environ Toxicol Chem 25:2177–2186. 18. Sørmo EG, Salmer MP, Jenssen B, Hop H, Baek K, Kovacs KM, Lydersen C, Falk-Petersen S, Gabrielsen GW, Lie E, Skaare JU. 2006. Biomagnification of polybrominated diphenyl ether and hexabromocyclododecane flame retardants in the polar bear food chain in Svalbard, Norway. Environ Toxicol Chem 25:2502–2511. 19. Tomy GT, Budakowski W, Halldorson T, Whittle DM, Keir M, Marvin C, Macinnis G, Alaee M. 2004. Biomagnification of a- and ghexabromocyclododecane isomers in a Lake Ontario food web. Environ Sci Technol 38:2298–2303. 20. J€ orundsd ottir H, L€ ofstrand K, Svavarsson J, Bignert A, Bergman Å . 2013. Polybrominated diphenyl ethers (PBDEs) and hexabromocyclododecane (HBCD) in seven different marine bird species from Iceland. Chemosphere 93:1526–1532. 21. Shaw SD, Berger ML, Brenner D, Kannan K, Lohmann N, P€apke O. 2009. Bioaccumulation of polybrominated diphenyl ethers and hexabromocyclododecane in the northwest Atlantic marine food web. Sci Total Environ 407:3323–3329. 22. Wu MH, Zhu JY, Tang L, Liu N, Peng BQ, Sun R, Xu G. 2014. Hexabromocyclododecanes in surface sediments from Shanghai, China: Spatial distribution, seasonal variation and diastereoisomerspecific profiles. Chemosphere 111:304–311. 23. Haukås M, Hylland K, Berge JA, Nygård T, Marius E. 2009. Spatial diastereomer patterns of hexabromocyclododecane (HBCD) in a Norwegian fjord. Sci Total Environ 407:5907–5913. 24. K€ oppen R, Becker R, Esslinger S, Nehls I. 2010. Enantiomer-specific analysis of hexabromocyclododecane in fish from Etnefjorden (Norway). Chemosphere 80:1241–1245. 25. Covaci A, Gerecke AC, Law RJ, Voorspoels S, Kohler M, Heeb NV, Leslie H, Allchin CR, de Boer J. 2006. Hexabromocyclododecanes (HBCDs) in the environment and humans: A review. Environ Sci Technol 40:3679–3688. 26. Bignert A, Danielsson S, Faxneld S, Miller A, Nyberg E, Berger U, Borg H, Eriksson U, Holm K, Nylund K, Egeb€ack A-L., Haglund P. 2013. Comments Concerning the National Swedish Contaminant Monitoring Programme in Marine Biota, 2013. Swedish Museum of Natural History, Stockholm. 27. Ciutat A, Widdows J, Readman JW. 2006. Influence of cockle Cerastoderma edule bioturbation and tidal-current cycles on resuspension of sediment and polycyclic aromatic hydrocarbons. Marine Ecology Progress Series 328:51–64. 28. Hedman JE, Bradshaw C, Thorsson MH, Gilek M, Gunnarsson JS. 2008. Fate of contaminants in Baltic Sea sediments: The role of bioturbation and settling organic matter. Marine Ecology Progress Series 356:25–38. 29. BACC (BALTEX Assessment of Climate Change). 2008. Assessment of Climate Change for the Baltic Sea Basin. Springer-Verlag, Berlin Heidelberg, Germany. 30. Sellstr€ om U, Kierkegaard A, de Wit C, Jansson B. 1998. Polybrominated diphenyl ethers and hexabromocyclododecane in sediment and fish from a Swedish river. Environ Toxicol Chem 17:1065–1072. 31. Kuiper RV, Cant on RF, Leonards PEG, Jenssen BM, Dubbeldam M, Wester PW, van den Berg M, Vos JG, Vethaak AD. 2007. Long-term exposure of European flounder (Platichthys flesus) to the flameretardants tetrabromobisphenol A (TBBPA) and hexabromocyclododecane (HBCD). Ecotoxicol Environ Safety 67:349–360.

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Environ Toxicol Chem 9999, 2015

32. Gunnarsson J, Bj€ork M, Gilek M, Granberg M, Rosenberg R. 2000. Effects of eutrophication on contaminant cycling in marine benthic systems. Ambio 29:252–259. 33. Fjeld E, Schlabach M, Berge JA, Green N, Eggen T, Snilsberg P, Vogelsang C, Rognerud S, Kjellberg G, Enge EK, Dye C, Gundersen H. 2005. Screening of selected new organic contaminants 2004. Brominated flame retardants, perfluorinated alkylated substances, irgarol, diuron, BHT and dicofol. Report TA 2096, NIVA, Norway (in Norwegian). 34. Smolarz K, Berger A. 2009. Long-term toxicity of hexabromocyclododecane (HBCDD) to the benthic clam Macoma balthica (L.) from the Baltic Sea. Aquat Toxicol 95:239–247. 35. Bonsdorff E, Blomqvist EAM. 1993. Biotic couplings on shallow water soft bottoms: Examples from the northern Baltic Sea. Oceanography and Marine Biology: An Annual Review 31:153–176. 36. Du M, Lin L, Yan C, Zhang X. 2012. Diastereoisomer- and enantiomerspecific accumulation, depuration, and bioisomerization of hexabromocyclododecanes in zebrafish (Danio rerio). Environ Sci Technol 46:1104011046.

C. Bradshaw et al. 37. Su G, Saunders D, Yu Y, Yu H, Zhang X, Liu H, Giesy JP. 2014. Occurrence of additive brominated flame retardants in aquatic organisms from Tai Lake and Yangtze River in Eastern China, 20092012. Chemosphere 114:340–346. 38. Zhang Y, Sun H, Ruan Y. 2014. Enantiomer-specific accumulation, depuration, metabolization andisomerization of hexabromocyclododecane (HBCD) diastereomers in mirror carp from water. J Hazard Mater 264:8–15. 39. European Commission. 2014. Study for the Review of the List of Restricted Substances Under RoHS2. Final Report. Annex V: ROHS Annex II Dossier for HBCDD. Reference: ENV.C.2/ETU/2012/0021. 40. European Chemicals Bureau. 2003. Technical Guidance Document on Risk Assessment in support of Commission Directive 93/67/EEC on Risk Assessment for new notified substances, Commission Regulation (EC) No 1488/94 on Risk Assessment for existing substances, Directive 98/8/EC of the European Parliament and of the Council concerning the placing of biocidal products on the market. Part II. European Commission.

Effects of benthos, temperature, and dose on the fate of hexabromocyclododecane in experimental coastal ecosystems.

The authors studied the fate of the brominated flame retardant hexabromocyclododecane (HBCDD) added in a particulate suspension to experimental ecosys...
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