From Sediment to Top Predators: Broad Exposure of Polyhalogenated

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From Sediment to Top Predators: Broad Exposure of Polyhalogenated Carbazoles in San Francisco Bay (U.S.A.) Yan Wu, Hongli Tan, Rebecca Sutton, and Da Chen Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b05733 • Publication Date (Web): 23 Jan 2017 Downloaded from http://pubs.acs.org on January 23, 2017

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Environmental Science & Technology

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From Sediment to Top Predators: Broad Exposure of Polyhalogenated

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Carbazoles in San Francisco Bay (U.S.A.)

3 Yan Wu, † Hongli Tan, ‡,† Rebecca Sutton, § Da Chen ‡,†,*

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†

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Illinois University, Carbondale, Illinois 62901, United States

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‡

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Health, and Guangdong Key Laboratory of Environmental Pollution and Health, Jinan

Cooperative Wildlife Research Laboratory and Department of Zoology, Southern

School of Environment, Guangzhou Key Laboratory of Environmental Exposure and

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University, Guangzhou, 510632, China

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§

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United States

San Francisco Estuary Institute, 4911 Central Avenue, Richmond, California 94804,

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*Corresponding author email: [email protected]; phone: (618) 453-6946; fax: (618) 453-

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6944.

ACS Paragon Plus Environment

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Abstract. The present study provides the first comprehensive investigation of

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polyhalogenated carbazoles (PHCZ) contamination in an aquatic ecosystem. PHCZs have

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been found in soil and aquatic sediment from several different regions, but knowledge of

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their bioaccumulation and trophodynamics is extremely scarce. This work investigated a

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suite of 11 PHCZ congeners in San Francisco Bay (United States) sediment and

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organisms, including bivalves (n = 6 composites), sport fish (n = 12 composites), harbor

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seal blubber (n = 18), and bird eggs (n = 8 composites). The most detectable congeners

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included 3,6-dichlorocarbazole (36-CCZ), 3,6-dibromocarbazole (36-BCZ), 1,3,6-

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tribromocarbazole (136-BCZ), 1,3,6,8-tetrabromocarbazole (1368-BCZ), and 1,8-

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dibromo-3,6-dichlorocarbazole (18-B-36-CCZ). The median concentrations of ΣPHCZs

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were 9.3 ng/g dry weight in sediment and ranged from 33.7 to 164 ng/g lipid weight in

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various species. Biomagnification was observed from fish to harbor seal and was mainly

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driven by chlorinated carbazoles, particularly 36-CCZ. Congener compositions of PHCZs

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differed among species, suggesting that individual congeners may be subject to different

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bioaccumulation or metabolism in species occupying various trophic levels in the studied

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aquatic system. Toxic equivalent (TEQ) values of PHCZs were determined based on their

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relative effect potencies (REP) compared to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD).

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The median TEQ was 1.2 pg TEQ/g dry weight in sediment and 4.8 – 19.5 pg TEQ/g

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lipid weight in biological tissues. Our study demonstrated the broad exposure of PHCZs

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in San Francisco Bay and their characteristics of bioaccumulation and biomagnification

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along with dioxin-like effects. These findings raise the need for additional research to

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better elucidate their sources, environmental behavior, and fate in global environments.

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TOC Art

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INTRODUCTION

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Marine and freshwater systems are subject to contamination by a large variety of

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halogenated substances. Well-known halogenated contaminants that have been

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extensively investigated worldwide, such as polybrominated diphenyl ethers (PBDEs)

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and polychlorinated biphenyls (PCBs) as well as many others, have been added to the

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persistent organic pollutants (POPs) list of the Stockholm Convention, due to their global

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contamination and toxic potencies.1-4 Recently, a number of polyhalogenated carbazoles

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(PHCZs) were discovered in soil and aquatic systems from different regions. Chlorinated

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and brominated carbazoles, including 3-chlorocarbazole (3-CCZ), 3-bromocarbazole (3-

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BCZ), 3,6-dibromocarbazole (36-BCZ), 1,3,6,8-tetrabromocarbazole (1368-CCZ) or

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1,3,6,8-tetrabromocarbazole (1368-BCZ), were reported in sediment from the Laurentian

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Great Lakes of North America, the Saginaw River system (United States or U.S.), the

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Ontario River (Canada), the North Sea Estuary (Germany), the Lippe River (Germany)

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and Lake Tai (China), as well as in soil from Germany and Greece.5-15 Carbazoles

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substituted with a combination of chlorine and bromine atoms (e.g., 1,8-dibromo-3,6-

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dichlorocarbazole or 18-B-36-CCZ) were also found in sediment from Southern Ontario,

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Canada and the Great Lakes.5,6,9 The Great Lake studies identified a number of other

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PHCZs with various combinations of halogen (bromine, chlorine, and iodine)

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substitutions, including Br2-, Br3-, Br4-, Br3Cl-, Br3ClI-, Br4Cl-, Br4I-, and Br5-

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carbazole.5,6 To date, the potential sources of the variety of PHCZ substances have not

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been thoroughly studied. Halogenated indigo dyes have been suggested as potential

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sources of 1368-BCZ and 18-B-36-CCZ, as well as some other PHCZs.16 These PHCZs

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may be formed as impurities in the final products of halogenated indigo dye formulations.


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Natural origins have also been suggested for selected PHCZs or their precursors.5,17

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Selected PHCZs can be produced through enzymatic synthesis of bromo- and chloro-

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carbazoles by chloroperoxidase from marine fungus Caldariomyces fumago in water.18

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But the extreme scarcity in relevant information and studies made it impossible to

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elucidate the sources of PHCZs and the volume of anthropogenic input.

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In addition to the discovery of PHCZs in environmental compartments, recent

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studies also investigated toxicological properties of selected halogenated carbazoles,

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including dioxin-like effects, and developmental, cardiotoxic, and mutagenic activities.19-

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injection dose of 30 or 60 mg/kg body weight (bw) for five consecutive days or at a

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single injection of 50 – 300 mg/kg bw resulted in the induction of a dominant lethal

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mutation or abnormal sperm heads, respectively.22 In zebrafish (Danio rerio), chronic

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exposure to both 27-BCZ and 2367-CCZ resulted in developmental toxicity in embryos at

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nano-molar levels and induced the expression of cytochrome P450 enzyme CYP1A1 in

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the heart area at micro-molar levels.20 The aryl hydrocarbon receptor (AhR)-mediated

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effects in MDA-MD-468 human breast cancer cells were also reported for 3-CCZ, 36-

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CCZ, 27-BCZ, 2367-CCZ and 18-B-36-CCZ, as well as some other PHCZs.19

Treatment of adult male mice (Mus musculus) with carbazole at a daily intraperitoneal

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Despite an increasing number of reports on the environmental occurrence and

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toxicity of PHCZs, knowledge of their bioavailability and consequent contamination in

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aquatic and terrestrial organisms remains extremely limited. The present work introduces

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the first comprehensive study to address the contamination of PHCZs in abiotic and

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biotic components of one of the most studied estuarine systems in North America, San

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Francisco Bay (U.S.). San Francisco Bay has been the subject of long-term monitoring of


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anthropogenic pollution, which revealed some of the highest concentrations of POPs

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(e.g., PBDEs and PCBs) worldwide in its sediment and biota, mainly due to a dense

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urban population and associated activities.23-27 The specific objectives of this study were

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to: (1) investigate spatial distributions of PHCZs in San Francisco Bay sediment; (2)

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evaluate the bioaccumulation and biomagnification potency of PHCZs; and, (3)

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fingerprint the composition patterns of PHCZ congeners in sediment and biota.

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MATERIALS AND METHODS

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Chemicals and Reagents. The reference standards of 3-CCZ, 36-CCZ, 1368-CCZ, 2367-

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CCZ, 1,3,6-tribromocarbazole (136-BCZ), 1-bromo-3,6-dichlorocarbazole (1-B-36-

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CCZ), and 18-B-36-CCZ were purchased from Wellington Laboratories (Guelph, ON,

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Canada). The standards of 3-BCZ, 27-BCZ, and 36-BCZ were purchased from Sigma-

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Aldrich (St. Louis, Missouri). The reference standard of carbazole and 1368-BCZ were

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purchased from AccuStandard (New Haven, CT) and Florida Center for Heterocyclic

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Compounds of the University of Florida (Gainesville, FL), respectively. Surrogate

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standards 4'-fluoro-2,3',4,6-tetrabromodiphenyl ether (F-BDE69), 4'-Fluoro-2,3,3',4,5,6-

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hexabromodiphenyl ether (F-BDE160) and 2,2',3,4,4',5,6,6'-octachorobiphenyl (PCB-

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204), as well as internal standards 3'-Fluoro-2,2',4,4',5,6'-hexabromodiphenyl ether (F-

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BDE154) and decachlorodiphenyl ether (DCDE), were purchased from AccuStandard

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(New Haven, CT). A total of 20 PBDE congeners, including BDE-28, -47, -49, -66, -85,

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-99, -100, -138, -153, -154, -183, -196, -197, -201, -202, -203, -206, -207, -208 and -209,

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were purchased from AccuStandard. Diatomaceous earth and sodium sulfate (10-60

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mesh) were purchased from Fisher Scientific (Hanover Park, IL) and treated in a muffle


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furnace at 600 °C overnight (> 12 h) prior to use. Copper (50 mesh, granular, reagent

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grade), as well as high performance liquid chromatography (HPLC) grade solvents, was

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purchased from Fisher Scientific. Isolute® silica sorbent (average pore size: 60) was

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purchased from Biotage Inc. (Charlotte, NC, USA) and baked at 130 °C prior to use.

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Samples. Surficial sediment was collected at a depth of 0 to 5 cm using a Young-

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modified Van Veen grab with a surface area of 0.1 m2 at 26 different sites in San

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Francisco Bay in 2014 (Figure 1A; Table S2).23 Sediment was kept in pre-cleaned I-

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Chem jars after collection. A field blank was also prepared along with sediment

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collection. Transplanted bivalves (Mytilus californianus) were collected from a non-

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urban reference site (Tomales Bay; 38°18'00.0"N, 123°04'12.0"W) and deployed at six

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stations in the Bay for 90 days during the summer of 2014 (Figure 1B; Table S2).

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Approximately 30-40 bivalves were sampled from each site and their soft tissues were

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homogenized. Four species of sport fish, including white croaker (Genyonemus lineatus),

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striped bass (Morone saxatilis), jacksmelt (Atherinopsis californiensis), and white

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sturgeon (Acipenser transmontanus), were collected at 8 recreational fishing areas across

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the Bay in 2014 and 2009 (Figure 1B; Table S2). Tissues from 1 to 21 individual fish of

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each species from the same site and collection year were homogenized to make a

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composite. Jacksmelt were processed whole but with the head, tail, and viscera removed.

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White croaker were processed whole, and all other samples were processed as muscle

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fillets with skin removed. A total of 12 fish composites were processed for chemical

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analysis. Eggs of double-crested cormorants (Phalacrocorax auritus) were collected in

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2016 at three locations within the Bay: Wheeler Island (North Bay), Richmond Bridge

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(Central Bay), and Don Edwards Wildlife Refuge (South Bay) (Figure 1B; Table S2).


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Two to three egg composites, containing 7 – 10 eggs per composite, were available from

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each site, resulting in a total of eight composites for chemical analysis. Blubber tissues

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were sampled in 2007 – 2015 from 18 adult or sub-adult harbor seals residing in the Bay

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(Figure 1B; Table S2). These harbor seals had various conditions during blubber

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collection, i.e. captured alive (n=5), deceased during treatment (n=3), euthanized (n=3),

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or deceased in the wild (n=7). Detailed blubber collection method is provided in the

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Supporting Information. All samples were shipped with ice packs to the analytical

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laboratory at the Southern Illinois University and stored at -20 °C prior to chemical

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analysis.

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Sample Preparation. Sample pretreatment and cleanup procedures described

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below applied to the determination of both PHCZ and PBDE congeners. Sediment

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samples were freeze-dried for 48 hours and sieved through a 100-micron stainless cloth

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sieve (Hogentogler & Co. Inc., Columbia, MD). Dry sediments were analyzed for the

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total organic carbon (TOC) percentages using a FlashEA® 1112 Nitrogen and Carbon

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Analyzer (Thermo Fisher Scientific, Waltham, MA). Bivalve composites were freeze-

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dried for 48 hours. Approximately 5 g of dried sediment, 1.0 g of bivalve and fish

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composites, 3.0 g of egg composites, or 0.11 – 0.49 g of blubber was ground with

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diatomaceous earth. After spiking with surrogate standards (FBDE-69, FBDE-160, and

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PCB-204; 50 ng each), the sample was subjected to accelerated solvent extraction

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(Dionex ASE 350, Sunnyvale, CA, USA) with dichloromethane (DCM) at 100 °C and

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1500 psi. Copper powder activated by concentrated hydrochloric acid was used to

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remove sulfur in the sediment extract.8 Tissue extracts were subjected to gravimetric

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determination of lipid content by using 10% of the extract. The extract after copper


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treatment or lipid determination was purified by a Shimadzu Prominence Semi-Prep

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HPLC (Shimadzu America Inc., Columbia, MD) equipped with a Phenogel gel

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permeation chromatography (GPC) column (300 × 21.2 mm, 5µ, 100Å; Phenomenex,

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Inc., Torrance, CA). The mobile solvent was 100% DCM and the flow rate was set at 4

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mL/min. Target compounds (including PHCZ and PBDE congeners) were collected in

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the fraction ranging from 16 to 45 minutes. The resulting fraction was further cleaned and

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separated on a 2-g Isolute® silica solid phase extraction (SPE) column packed into a 6 mL

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polyesters SPE tube (SiliCycle Inc., Quebec City, Canada). The SPE column was pre-

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washed with 10 mL hexane (HEX) to condition the silica gel sorbent. After the sample

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was loaded, the first fraction was eluted with 3 mL HEX and was discarded. The second

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fraction that contained target PHCZs was eluted with 11 mL of a mixture of HEX and

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DCM (40:60, v/v). The latter fraction was concentrated to approximately 200 µL and

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transferred to an insert tube in a gas chromatography (GC) vial. Internal standards FBDE-

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154 and DCDE (100 ng each) were added prior to instrumental analysis.

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Instrumental Analysis. The separation and quantification of the target PHCZs, as

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well as the surrogate and internal standards, was performed on an Agilent 7890B gas

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chromatography (GC; Agilent Technologies, Palo Alto, CA) coupled to a single

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quadrupole mass analyzer (Agilent 5977A MS) in either electron-capture negative

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ionization (ECNI) or electron impact (EI) mode. Our previous study has revealed that 3-

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CCZ, 36-CCZ, or 3-BCZ exhibited a greater sensitivity in EI versus ECNI mode, whereas

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the other PHCZ congeners had a much better sensitivity in ECNI mode.12 The column

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used for GC-MS analysis was a 30 m HP-5MS column (0.25 mm i.d., 0.25 µm, J&W

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Scientific, Agilent Tech.). The injector was operated in pulsed-splitless mode, held at


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260 °C. Initial oven temperature was held at 50 °C for 3 min, increased to 150 °C at 10

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°C/min, and then ramp to 300 °C at 5 °C/min and held for 10 min. The quantification and

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confirmation of each target compound was achieved via selected ion monitoring (SIM)

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for its characteristic ions under ECNI or EI mode (Table S1). Instrumental analysis of

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PBDE congeners was described in detail in Supporting Information.

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Quality Assurance and Control. The QA/QC practices in the present study

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included the analysis of field or laboratory procedural blanks, spiking recovery tests, and

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replicates of randomly chosen authentic samples, as well as monitoring the recoveries of

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surrogate standards. No PHZCs were detected in any field or laboratory blanks. Analyses

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of reference sediment (collected from Au Sable River, Michigan, USA), fish (Tilapia,

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Oreochromis niloticus) fillets, and chicken (Gallus domesticus) eggs spiked with known

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amounts of PHCZs (25 ng of each congener) revealed a mean recovery of individual

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PHCZ congeners ranging from 71.6% to 113.8% (adjusted with surrogate recoveries) in

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all matrices. Fish fillets and chicken eggs were purchased from a local supermarket and

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were confirmed to be free of PHCZs prior to spiking experiments. No PHCZs were

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quantifiable in reference sediment.

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revealed the relative standard deviations (RSDs) less than 6.7% for the summed

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concentrations of detected PHCZ congeners (referred to ΣPHCZs) and ranged from 4.8%

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to 9.1% for individual congeners. Recoveries of surrogate standards were 92.3 ± 13.5%

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(mean ± standard deviation) for FBDE-160 and 88.8 ± 11.4% for PCB-204, respectively.

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The method limits of quantification (MLOQs) were assessed by multiplying a Student’s

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t-value designated for a 99% conference level with standard deviations in the replicate

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analyses (n = 8) of reference sediment, fish, or eggs, following the protocol introduced in

Replicate analysis of random samples (n = 3)


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Wu et al.8 The MLOQs ranged from 0.1 to 0.3 ng/g dw in sediment, 0.8 – 1.5 ng/g lw in

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fish, and 0.5 – 1.2 ng/g lw in bird eggs (Table S1). The MLOQs in bivalve and harbor

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seal were adjusted from those determined in fish with relative lipid contents. Analytes

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with instrumental responses below the instrumental detection limit (IDL), defined as a

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concentration yielding a signal-to-noise ratio (S/N) of 5, were considered non-detectable

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(nd). For measurements below MLOQs or non-detectable for an analyte with detection

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frequency above 60%, a regression plotting method was applied to assign values for

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statistical analysis.28

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Data Analysis. The PHCZ concentrations in San Francisco samples were

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adjusted with recoveries of surrogate standards (i.e. FBDE-160 for ECNI and PCB-204

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for EI analyses) and reported as ng/g dry weight (dw) for sediment and ng/g lipid weight

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(lw) for biological samples. The suitability of using these surrogate standards for

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recovery adjustment has been demonstrated in previous studies.8,13 The Student’s t-test

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and the Pearson’s correlative analyses, as well as the principal component analysis

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(PCA), were conducted using the OriginPro 9.0 (OriginLab Corporation). The analysis of

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variance (ANOVA) was conducted using the PASW Statistics 18.0 (IBM Inc.). The level

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of significance is set as α = 0.05. Concentrations of PBDEs in all samples of the present

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study have not been published elsewhere.

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RESULTS AND DISCUSSION

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PHCZs in San Francisco Bay Sediment. PHCZs were detected in sediment samples

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collected from all 26 studied sites, revealing their wide distribution in the Bay (Figure

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1A). The concentrations of ΣPHCZs ranged from 1.7 to 20.5 ng/g dw (Table S3). PHCZs


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were also reported in other aquatic systems,5-13 but the number of investigations to date

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remains overall very limited. The median concentration (i.e., 9.3 ng/g dw) of ΣPHCZs in

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the Bay was lower than what has been reported in surface sediment from the Saginaw

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River basin (Michigan, U.S.; median: 18.9 ng/g dw) and the Great Lakes (median: 38.0

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ng/g dw), but much greater than the levels reported in sediment collected from rivers and

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coastal water of the North Sea estuary (Germany; median: 0.23 ng/g dw) and Lake Tai of

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China (median: 1.5 ng/g dw).5,8,10,13 It should be noted that the Great Lakes study

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included a number of additional congeners compared to those determined in the present

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study, which do not have corresponding commercial reference standards available and

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were subjected to semi-quantification.5 Selected PHCZ congeners were also identified,

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quantified, or semi-quantified in sediment from the Ontario River (Canada) and Lake

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Michigan (U.S.) sediment cores.6,7,9 Their detection in various watersheds may suggest a

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broad distribution of these chemicals in global aquatic ecosystems.

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Spatial distribution of PHCZs revealed some elevated concentrations at sites

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along the eastern side of the Central and South Bay, agreeing with previous findings for

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PBDEs and PCBs at these sites.23,24 These elevated concentrations were likely due to

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greater urbanization as well as geographic and hydrological features. The South Bay

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extends from the Central Bay toward San Jose, while the North Bay extends from the

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freshwaters at the mouths of the Sacramento and San Joaquin rivers through the saline

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waters of Suisun and San Pablo Bays.25 Central and South Bay waterfronts include the

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majority of the industrial, port, and shipping activities in the region.26 Highly populated

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metropolitan centers including San Francisco and Oakland are located in the Central Bay,

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while San Jose is located in the South Bay. The Central and South Bay segments in total


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receive approximately 76% of the region’s wastewater inflow into the Bay.25,27 Given

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that some PHCZs were suggested as impurities in the final products of historically

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manufactured halogenated indigo dye formulations,16 they may be present in industrial or

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municipal wastewater. Hydrological features may also be important. Surface waters in

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the South Bay experience the least amount of mixing with non-wastewater effluent flow,

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particularly in the dry seasons, and thus have higher hydraulic residence time relative to

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other Bay segments.26 By contrast, the northern Bay segment receives up to 90% of the

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Estuary’s freshwater inflow;25 while these river discharges are influenced by upstream

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urban contaminants, they may nevertheless dilute environmental contaminants

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concentrated in local urban runoff or wastewater effluent and result in short residence

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time.

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However, statistical analysis did not reveal a significant difference in dry weight

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(p = 0.13) or TOC (p = 0.35) based ΣPHCZ concentrations among the North, Central, and

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South Bay segments, although a north-to-south increasing gradient has been reported for

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BDE-209 and PCBs.23,24 The TOC of sediment significantly correlated with ΣPHCZ

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concentrations (p < 0.001), but did not differ significantly among the Bay segments. Of

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note, a few sites from the northern segment exhibited sediment concentrations above the

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median value (Figure 1A). Some of the northern sites were also reported to have elevated

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PBDE concentrations in sediment.23 Possible anthropogenic sources or natural origins

292

may be present in the North Bay, but current information is insufficient to draw any

293

conclusion.

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PHCZs Are Bioavailable and Biomagnify.

PHCZs were detected in all

biological samples, including transplanted bivalves, sport fish, cormorant eggs, and


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harbor seal blubber (Figure 2; Table S3). Bivalves such as mussels are stationary

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organisms that filter particles from water. Contaminant concentrations in bivalve tissues

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are not only a good indicator of local contamination, but also an ideal measurement of the

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bioavailability of environmental pollutants because bivalves generally perform little

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metabolic transformation.29 Concentrations of ΣPHCZs in bivalve composites from the

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six sites within the Bay ranged from 8.3 to 76.1 ng/g lw, with a median concentration of

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33.7 ng/g lw. These levels were greater than that in the reference site (6.8 ng/g lw)

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located in the Tomales Bay, indicating elevated contamination in the estuarine system

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subject to urban influences. It is also noted that bivalves studied in the present work were

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transplanted to the designated sites, thus subject to shorter exposure time than resident

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bivalves, which may be exposed at a greater level.

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PHCZs were also present in sport fish, cormorant eggs, and harbor seal blubber, at

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median concentrations of 53, 155, and 164 ng/g lw, respectively. The occurrence in

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various species occupying different trophic levels, from benthic invertebrates to top

310

predators, demonstrates substantial bioavailability of PHCZs and their broad exposure in

311

the Bay ecosystem. The estimated bioconcentration factors (BCFs) of PHCZs via the

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U.S. Environmental Protection Agency (EPA) Estimation Program Interface (EPI) Suite

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Version 4.11 have logarithmically transformed values ranging from 2.73 to 3.15 (Table

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S1), except for 3-CCZ (2.43) and 3-BCZ (2.47). These estimated values are below the

315

criteria for bioaccumulation (i.e. log BCF = 3.3) recommended by the European

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Commission Registration, Evaluation and Authorization of Chemicals (REACH)

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program.30 However, studies have suggested that a conservative interval of 0.75 log units

318

should be considered when employing the BCF criteria, given the variability of the


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experimental BCF data used for model predictions.31,32 Thus, a chemical with estimated

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log BCF less than 3.3 cannot be safely classified as not bioaccumulative. Indeed, the

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Arnot/Gobas bioaccumulation factor (BAF) model suggests that an organic chemical with

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log Kow (octanol-water partitioning coefficient) value > 4 may possess bioaccumulation

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potential in aquatic food webs.30 Given the lack of experimentally determined data, the

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EPA EPI program was used to calculate log Kow for the analyzed PHCZ congeners, which

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range from 4.12 to 6.85, except for 3-CCZ (calculated log Kow = 3.94). Therefore, both

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model predictions and our environmental findings suggest considerable bioavailability

327

for the majority, if not all, of PHCZ congeners.

328

The data suggest biomagnification of PHCZs from fish to harbor seal. Three of

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the four investigated sport fish species, including white croaker, striped bass, and

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jacksmelt, have been demonstrated to constitute a portion of the diet of Pacific harbor

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seals which are opportunistic feeders, even though they do not represent seals’ major

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food sources.33 The biomagnification factor (BMF) of ΣPHCZs, calculated as the ratio of

333

median ΣPHCZ concentration in seal blubber to that in fish composites of each species,

334

was determined to be 6.1, 3.6 and 2.7 from white croaker, jacksmelt or striped bass to

335

seal, respectively. The BMFs for ΣPHCZs were even greater than those of ΣPBDEs in the

336

same fish-seal food chains (i.e., 1.0 – 2.9). The BMFs were also determined for

337

individual PHCZ congeners (Table S4). The results indicated that the biomagnification of

338

PHCZs was mainly driven by chlorinated carbazoles, particularly 36-CCZ which

339

exhibited statistically greater BMFs (3.3 – 7.5) than other congeners (i.e., 1368-CCZ, 36-

340

BBZ and 136-BCZ; ANOVA with Fisher’s post-hoc test, F3,8 = 11.5, p = 0.003). The

341

marine mammal food web model developed by Kelly et al. suggests that


15


Page 16 of 35

342

biomagnification capacities of organic chemicals are primarily controlled by their

343

physicochemical properties, such as log Kow and log Koa, assuming no metabolic

344

transformation.34 The chemicals that are subject to great levels of biomagnification in

345

marine mammalian food webs normally have a log Kow ranging from ~4 to ~7.8 and a log

346

Koa > 7.34

347

biomagnification potential in marine mammalian food webs. However, most of the BMF

348

values determined for brominated carbazoles were less than one in our study, likely

349

suggesting elevated metabolism of these congeners in harbor seals. But additional studies

350

are required to support this hypothesis. We would also like to point out that the

351

abovementioned approach used to determine BMFs has various sources of uncertainty

352

and may be affected by a number of factors,35 including the non-dominance of the

353

investigated fish species in seals’ diet, variations in age, sex, or collection year of the

354

analyzed seal samples or variations in the species and collection year of fish composites,

355

as well as relatively small sample sizes for each species. These sources of uncertainty

356

may account for a large variability of the measured BMF values between studies. Indeed,

357

the measured BMFs for ΣPBDEs in the present study (1.0 – 2.9) were generally lower

358

than the previously reported values (i.e. 2 – 76) in fish – seal food chains from the

359

northwestern Atlantic, North Sea, and Svalbard (Norway).36-41 Nevertheless, a direct

360

comparison of the BMF values between PHCZs and PBDEs in the same food chains in

361

the present study may adequately demonstrate the biomagnification potency of PHCZs,

362

mostly driven by chlorinated congeners, in the studied ecosystem. Future studies are

363

needed to better elucidate the biomagnification or metabolism of individual PHCZ

364

congeners through prey-predator food chains or different trophic levels of a food web.

Most PHCZs meet these criteria and therefore may possess substantial


16

Page 17 of 35

365


Congener-Specific

Bioaccumulation

in

Different

Species.

Congener

366

compositions of PHCZs differed among studied species and sediment (Figure 3). The

367

PCA analysis of congener compositions revealed three clusters in the biplot (Figure 4).

368

Congener compositions of bivalves resembled closely that of sediment (Cluster I), which

369

were dominated by 36-CCZ, followed by 36-BCZ, but had elevated contributions by

370

1368-BCZ compared to other species. Fish and harbor seal (Cluster II) shared a similar

371

congener composition profile, which exhibits a dominance of 36-CCZ, followed by 136-

372

BCZ and then 36-BCZ. Cormorant eggs (Cluster III) differed from all other species in

373

composition profiles, and were dominated by 136-BCZ, followed by 36-CCZ and 36-

374

BCZ. The elevated composition of 136-BCZ may indicate its greater biomagnification

375

potential in piscivorous bird food chains compared to that in marine mammal food chains

376

or greater metabolism by harbor seal than cormorant. The differences revealed by PCA

377

may overall suggest congener-specific bioaccumulation, biomagnification, or metabolism

378

among benthic species (Cluster I), fish and mammals (Cluster II), and avian species

379

(Cluster III) in Bay food webs.

380

The dominance of 36-CCZ in most species and sediment from the Bay was

381

consistent with previous findings in sediment from some other regions.8,10,13 This

382

consistency may indicate a relative abundance of this specific PHCZ congener during the

383

synthesis and manufacturing or its stability during degradation. These characteristics may

384

be attributed to an electrophilic aromatic substitution pattern which favors halogen

385

substitution at the ortho and para positions relative to nitrogen on the carbazole, as well

386

as the enhanced stability of para-substituted products resulting from the high charge

387

density at the para positions compared to the ortho positions and the effect of the lone


17


Page 18 of 35

388

pair of electrons in the NH2 moiety.18,42,43 Indeed, a dominance of products with halide

389

substitution at the 3,6- positions has been found in enzymatic synthesis of PHCZs from a

390

source material of carbazole.18 The 36-BCZ was also frequently detected in the Bay, and

391

was the second dominant congener in bivalves and sediment.

392

Although cormorant eggs and harbor seal blubber revealed comparable

393

concentrations of ΣPHCZs, their composition profiles differed significantly, mainly due

394

to a dominance of 136-BCZ over other congeners in the cormorants. Although additional

395

evidence is needed, our data suggest congener-specific bioaccumulation/biomagnification

396

potential or metabolic differences between marine mammal and avian food webs in the

397

studied system. The effect of depuration kinetics (e.g., maternal transfer in mammals and

398

birds) on composition profiles cannot be ignored, but it needs further investigations.

399

Unfortunately, comparison with other studies regarding the accumulation pattern of

400

PHCZs is not yet possible due to the extreme scarcity of relevant knowledge. While our

401

work provides insights in the bioaccumulation of PHCZs in aquatic organisms, additional

402

efforts are essential to better elucidate the trophodynamics and fate of PHCZs in different

403

biological systems.

404

Implications and Future Research Needs. To better understand the abundance

405

of PHCZs relative to other important anthropogenic pollutants in San Francisco Bay, we

406

compared their concentrations in sediment and biological tissues with three groups of

407

POP substances: PBDEs, PCBs and hexachlorohexanes (HCHs). These contaminants

408

have been broadly distributed in global ecosystems and subjected to long-term

409

environmental monitoring in many important aquatic ecosystems, including San

410

Francisco Bay. Concentrations of ΣPHCZs were significantly greater (p < 0.001) than


18

Page 19 of 35


411

that of ∑PBDEs (summed concentration of detected PBDE congeners) in the same

412

sediment samples (0.2 – 5.8 ng/g dw) (Figure S1, Supporting Information). The median

413

concentration ratio of ΣPHCZs to ∑PBDEs was 5.2 in sediment, while the ratios ranged

414

from 0.10 to 0.29 in organisms examined in the present study. Significant correlations

415

(Figure 5) were also found between ∑PHCZ and ∑PBDE concentrations in the sediment

416

and biota (p < 0.01 in all species except for transplanted bivalves), suggesting that these

417

two groups of contaminants may share some common sources (e.g., wastewater effluent

418

and urban runoff) and possess similar environmental processes.

419

Concentrations of PHCZs were greater than those of ΣHCHs and comparable to

420

those of ΣPCBs in the sediment of San Francisco Bay (Figure S1). 44,45 Data of sediment-

421

associated PCBs and HCHs included in Figure S1 were from multiple sites, even though

422

the sites were not entirely consistent with those analyzed in the present study. The levels

423

of PHCZs were also greater than those of ΣHCHs in various biological species from the

424

Bay ecosystem, but were generally two to three orders of magnitude lower than ΣPCBs in

425

biota.44-46 It is noted that the number of PCB congeners under investigation differed

426

among studies, ranging from 46 to 209.44-46 The median concentration ratios of ΣPHCZs

427

to ∑PCBs and ΣPHCZs to ∑HCHs in the Bay organisms ranged from 0.004 to 0.07 and

428

1.7 – 10.9, respectively.44-46 These comparisons suggest PHCZs are among the groups of

429

halogenated contaminants with considerable abundances in the Bay aquatic ecosystems.

430

PBDE concentrations in the Bay have declined by approximately half in sport fish and

431

74-95% in bivalves and bird eggs, as a result of discontinuation in production since

432

2004.23 Concentrations of other legacy POPs (e.g., organochlorine pesticides) have also

433

declined over time in the Bay system.47 The future trends of PHCZ concentrations in Bay


19


Page 20 of 35

434

sediment and biota remain unknown due to the lack of temporal studies, thus warranting

435

further investigation.

436

Among a variety of potential toxicity activities of PHCZs, their dioxin-like

437

activity has raised particular concern.19-21,48 Riddell et al. reported a similarity between

438

PHCZs and other dioxin-like compounds in the structure-dependent induction of

439

cytochrome P450 (CYP450) 1A1 and CYP1B1 gene expression in aryl hydrocarbon-

440

responsive MDA-MB-468 breast cancer cells and that the magnitude of the induction

441

responses for the most active PHCZs (e.g., 2367-CCZ) was reported to be similar to that

442

for 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD).19 The relative effect potencies (REP) of

443

PHCZs (compared to TCDD) were estimated to range from 0.000013 to 0.00066 for

444

mono- to tetra-PHCZs.19 The REP can be used to represent toxic equivalency factors

445

(TEFs) of PHCZs, given that they were determined following an extensively used

446

approach for TEF estimations.49 Based on the REP measurements and using the equation

447

described in Wu et al. (also seen in Supporting Information),13 we estimated the median

448

toxic equivalent (TEQ) of PHCZs to be 1.2 pg TEQ/g dw in sediment and 4.8 – 19.5 pg

449

TEQ/g lw in biological tissues (Table S3). The TEQs of PHCZs appeared to be lower

450

than those reported for polychlorinated dibenzo-dioxins/furans (PCDDs/Fs) and co-planar

451

PCBs (i.e. 33.1 and 109 pg/g lw, respectively) in fish from California coastal waters

452

(collected in 2000-2001).50 Using the same approach, Wu et al. has estimated the TEQs

453

of PHCZs in Lake Tai (China) sediments ranging from 0 to 1.36 pg TEQ/g dw (median =

454

0.12).13 Pena-Abauurea et al. reported the TEQ for 18-B-36-CCZ alone ranging from

455

0.31 to 7.46 pg TEQ/g dw in stream sediment from Ontario (Canada),48 by employing a

456

REP of 0.0016 for this particular congener, different from what we used (0.0003) in the


20

Page 21 of 35


457

present study. Future studies are needed to unify the TEF values for the evaluation of

458

dioxin-like effects of PHCZs in global ecosystems.

459

In summary, this study has found that PHCZs are abundant in sediment of the San

460

Francisco Bay, and furthermore, they appear to have the ability to bioaccumulate and

461

biomagnify in marine mammal food webs. Bioaccumulation and biomagnification of

462

PHCZs may also be likely in terrestrial mammalian and human food webs according to

463

their physicochemical properties. These characteristics and existing evidence suggest the

464

PHCZs as a potential group of persistent organic pollutants with dioxin-like potency and

465

global distribution. However, knowledge gaps remain in many aspects of our current

466

understanding of these substances, including, but not limited to: (1) potential sources,

467

origins, or fate in aquatic systems; (2) occurrence and bioaccumulation in other

468

marine/freshwater ecosystems, non-aquatic systems and human living environments; (3)

469

rates of environmental degradation and metabolism; (4) long range transport potential;

470

and (5) other toxicities in addition to what has been reported. Thus, additional field and

471

laboratory-based research is critically needed in order to comprehensively assess the

472

environmental distribution, behavior, fate and effects of PHCZs in abiotic and biological

473

systems.

474 475

ACKNOWLEDGEMENT

476

The authors would like to acknowledge the San Francisco Estuary Institute and the

477

Regional Monitoring Program for Water Quality in San Francisco Bay (RMP) for

478

samples and financial support. We thank J. Davis, J. Sun, P. Trowbridge, and D. Yee for

479

assistance and constructive comments. We also thank D. Greig and B. Halaska for


21


Page 22 of 35

480

providing archived harbor seal blubber samples collected by The Marine Mammal Center

481

and California Academy of Sciences under stranding agreements from The Marine

482

Mammal Health and Response Program and J. Harvey, Moss Landing Marine

483

Laboratories, for wild-caught harbor seal samples collected under NMFS research permit

484

#16991. A. Rothert and S. Baer from SIU Core Facility for Ecological Analyses were

485

thanked for TOC analysis.

486 487

Supporting Information

488

Table S1-S4 and Figure S1 are available in the Supporting Information.

489 490

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(48) Pena-Abaurrea, M.; Robson, M.; Chaudhuri, S.; Riddell, N.; McCrindle, R.; Chittim,

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B.; Parette, R.; Jin, U.-H.; Safe, S.; Poirier, D.; Ruffolo, R.; Dyer, R.; Fletcher, R.;

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Helm, P.A.; Reiner, E.J. Environmental levels and toxicological potencies of a novel

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mixed halogenated carbazole. Emerg. Contam. 2016, 2, 166-172.

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(49) Van den Berg, M.; Birnbaum, L.; Bosveld, A.T.; Brunstrom, B.; Cook, P.; Feeley,

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M.; Giesy, J.P.; Hanberg, A.; Hasegawa, R.; Kennedy, S.W.; Kubiak, T.; Larsen,

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J.C.; van Leeuwen, F.X.; Liem, A.K.; Nolt, C.; Peterson, R.E.; Poellinger, L.; Safe,

651

S.; Schrenk, D.; Tillitt, D.; Tysklind, M.; Younes, M.; Waern, F.; Zacharewski, T.

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Toxic equivalency factors (TEFs) for PCBs, PCDDs, PCDFs for humans and

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wildlife. Environ. Health Perspect. 1998, 106, 775-792.

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PCDDs, PCDFs, and coplanar PCBs in edible fish from California coastal waters.

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Chemosphere 2006, 64, 276-28.


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657

FIGURE LEGENDS

658

FIGURE 1. Sampling sites of sediment (A) and biological samples (B), including

659

bivalve, sport fish, cormorant egg and harbor seal, in San Francisco Bay. The left bubble-

660

map (A) shows the concentrations (ng/g dry weight) of ∑PHCZs in sediment at each site.

661

FIGURE 2. Concentrations of ∑PHCZs in sediment (ng/g dry weight) and biological

662

samples (ng/g lipid weight) from San Francisco Bay. The concentration of the bivalve

663

composite from a reference site was excluded from the graph.

664

FIGURE 3. Congener compositions of PHCZs in sediment and biological samples. Error

665

bars represent standard deviations.

666

FIGURE 4. Biplot from the principle component analysis of PHCZ congener distribution

667

patters. S, B, E, F and H represent sediment, bivalve, cormorant egg, sport fish, and

668

harbor seal, respectively.

669

FIGURE 5. Correlations between concentrations of ∑PHCZs and ∑PBDEs in sediment (r

670

= 0.41, p = 0.021), sport fish (r = 0.95, p < 0.0001), cormorant egg (r = 0.94, p < 0.001),

671

and harbor seal blubber (r = 0.71, p < 0.0001).

672 673 674


30

Page 31 of 35


675

676 677

FIGURE 1

678 679 680


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Page 32 of 35

681

682 683

FIGURE 2

684 685 686 687 688 689 690 691


32

Page 33 of 35


692 693

FIGURE 3

694 695 696 697 698 699 700 701


33


Page 34 of 35

702 703

FIGURE 4

704 705 706 707 708 709 710 711 712


34

Page 35 of 35


713 714

FIGURE 5


35

From Sediment to Top Predators: Broad Exposure of Polyhalogenated

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