R E S E A R CH R E V I EW

Climate change–contaminant interactions in marine foodwebs: Toward a conceptual framework

Juan Jos�e Alava1,2 | William W. L. Cheung1 | Peter S. Ross2 | U. Rashid Sumaila1

1Global Fisheries Cluster, Institute for the

Oceans and Fisheries, University of British

Columbia, Vancouver, BC, Canada

2Ocean Pollution Research Program,

Coastal Ocean Research Institute,

Vancouver Aquarium Marine Science

Centre, Vancouver, BC, Canada

Correspondence

Juan Jos�e Alava, University of British

Columbia, Vancouver, BC, Canada.

Email: j.alava@oceans.ubc.ca

Funding information

Mitacs-SSHRC Joint Initiative; OceanCanada

Partnership; Coastal Ocean Research

Institute; Nippon Foundation-Nereus

Program; Natural Sciences and Engineering

Research Council of Canada (NSERC)

Abstract

Climate change is reshaping the way in which contaminants move through the glo-

bal environment, in large part by changing the chemistry of the oceans and affecting

the physiology, health, and feeding ecology of marine biota. Climate change-asso-

ciated impacts on structure and function of marine food webs, with consequent

changes in contaminant transport, fate, and effects, are likely to have significant

repercussions to those human populations that rely on fisheries resources for food,

recreation, or culture. Published studies on climate change–contaminant interactions

with a focus on food web bioaccumulation were systematically reviewed to explore

how climate change and ocean acidification may impact contaminant levels in mar-

ine food webs. We propose here a conceptual framework to illustrate the impacts

of climate change on contaminant accumulation in marine food webs, as well as the

downstream consequences for ecosystem goods and services. The potential impacts

on social and economic security for coastal communities that depend on fisheries

for food are discussed. Climate change–contaminant interactions may alter the

bioaccumulation of two priority contaminant classes: the fat-soluble persistent

organic pollutants (POPs), such as polychlorinated biphenyls (PCBs), as well as the

protein-binding methylmercury (MeHg). These interactions include phenomena

deemed to be either climate change dominant (i.e., climate change leads to an

increase in contaminant exposure) or contaminant dominant (i.e., contamination

leads to an increase in climate change susceptibility). We illustrate the pathways of

climate change–contaminant interactions using case studies in the Northeastern

Pacific Ocean. The important role of ecological and food web modeling to inform

decision-making in managing ecological and human health risks of chemical pollu-

tants contamination under climate change is also highlighted. Finally, we identify the

need to develop integrated policies that manage the ecological and socioeconomic

risk of greenhouse gases and marine pollutants.

K E YWORD S

acidification, climate change, contaminants, food web bioaccumulation, ocean warming, organic

mercury (MeHg), persistent organic pollutants (POPs), polychlorinated biphenyls (PCBs)

1 | INTRODUCTION

Marine pollution and anthropogenic climate change are two of the

dominant human-induced threats impacting the world’s oceans at

different scales in the Anthropocene era (Crutzen, 2002, 2006; Stef-

fen, Crutzen & McNeill, 2007). These two stressors are altering the

chemistry and health of our oceans and impacting the survival of

many marine species as well as the well-being of humans.

Received: 8 September 2016 | Revised: 7 February 2017 | Accepted: 8 February 2017

DOI: 10.1111/gcb.13667

3984 | © 2017 John Wiley & Sons Ltd wileyonlinelibrary.com/journal/gcb Glob Change Biol. 2017;23:3984–4001.

Ocean contamination by persistent organic pollutants (POPs) and

mercury is an issue of serious concern because these contaminants

are ubiquitous in the environment, detected at relatively high con-

centrations in some species, and driven by long-range atmospheric

transport from temperate, subtropical, and tropical areas to remote

parts of the world (e.g., Chiuchiolo, Dickhut, Cochran & Ducklow,

2004; Dickhut, Cincinelli, Cochran & Ducklow, 2005; Dietz, Outridge

& Hobson, 2009; Iwata, Tanabe, Sakai, Nishimura & Tatsukawa,

1994; Iwata, Tanabe, Sakai & Tatsukawa, 1993; Lamborg et al.,

2014; Streets et al., 2011; Wania, 2003; Wania & Mackay, 1993,

1996).

One class of POP of considerable concern has been the poly-

chlorinated biphenyls (PCBs), defined as persistent, bioaccumulative,

and toxic (i.e., PBT) within the context of the Stockholm Conven-

tions on POPs (Gobas, de Wolf, Verbruggen, Plotzke & Burkhard,

2009; UNEP, 2002). PCBs and other legacy POPs (e.g., dichloro-

diphenyl-trichloroethanes: DDTs) were largely phased out in the

1970s. However, some POPs are still used in developing countries

to control malaria vectors and pests, that is, organochlorine pesti-

cides such as DDTs (Alava, Ross, et al., 2011, Alava, Salazar et al.,

2011; Blus, 2003), and cycling of PCBs in the marine environment

lingers in many industrial areas (e.g., Blasius & Goodmanlowe, 2008;

Grant et al., 2011; Johannessen et al., 2008). POPs are bioaccumu-

lated by marine organisms and biomagnified in food webs, reaching

exposure concentrations above threshold effect levels in certain

populations of apex predators (e.g., Desforges et al., 2016; Kelly, Iko-

nomou, Blair, Morin & Gobas, 2007; Letcher et al., 2010; Ross, Ellis,

Ikonomou, Barrett-Lennard & Addison, 2000; Scheuhammer et al.,

2015). For example, a number of toxicological effects have been

attributed to PCBs in marine mammals, including molecular and cel-

lular alterations leading to immunotoxicity, endocrine disruption, and

reproductive impairment (e.g., Addison, 1989; Brouwer, Reijnders &

Koeman, 1989; Buckman et al., 2011; De Guise, Martineau, Beland

& Fournier, 1998; Desforges et al., 2016; Hall et al., 2006; Mos,

Cameron, Jeffries, Koop & Ross, 2010; Ross, De Swart, Van Loveren,

Osterhaus & Vos, 1996; Ross, De Swart, Timmerman, et al., 1996;

Tabuchi et al., 2006).

Mercury has become a metal of global concern due primarily to

health threats associated with biomagnification of its methylated

form, methylmercury (MeHg), in aquatic food webs (UNEP, 2013;

Wiener et al., 2007). Anthropogenic mercury is emitted to the atmo-

sphere and ocean mostly from metals mining (e.g., silver and gold),

coal-fired power plants, discharges from rivers, and hydroelectric

developments (i.e., flooding), where it can be carried long distances

to remote regions (e.g., Amos et al., 2014; Dietz et al., 2009; Dou-

glas, Amyot, Barkay, Berg & Ch�etelat, 2011; Krabbenhoft & Sunder-

land, 2013; Lamborg et al., 2014; Schartup et al., 2015; Streets

et al., 2011; UNEP, 2013). More than 80% of the mercury deposited

in the ocean is re-emitted to the atmosphere as gaseous mercury

(Hg0), driving the cycle of mercury through biogeochemical reservoirs

(Amos et al., 2014; Strode et al., 2007). Conversely, the total amount

of anthropogenic mercury present in the global oceans has been

estimated at 290 million moles, of which �66.67% reside in waters

<1000 m (Lamborg et al., 2014). As a consequence, mercury concen-

trations in the global ocean are expected to continue increasing

(Mason et al., 2012; Sunderland & Mason, 2007). Oxic surface

waters, the mesopelagic marine environment, and deep waters

appear to be sources of MeHg for the uptake and bioaccumulation

through the food web of top predators due to methylation processes

occurring in these zones (Kraepiel, Keller, Chin, Malcolm & Morel,

2003; Mason, Rolfhus & Fitzgerald, 1998; Peterson, Ackerman &

Costa, 2015; Schartup et al., 2015; Sunderland, Krabbenhoft, Mor-

eau, Strode & Landing, 2009). MeHg is highly neurotoxic and

nephrotoxic and bioaccumulates and biomagnifies throughout the

food web via dietary uptake, reaching the highest concentrations in

organisms at the top of the food web (Fort et al., 2015; Scheuham-

mer et al., 2015; Wiener et al., 2007). However, MeHg can be par-

ticularly toxic even at low concentrations (Dietz et al., 2013).

Consumption of mercury-contaminated fish is linked to disease in

humans, including the neurological disorders due to MeHg poisoning

(i.e., Minamata disease) (Ishikawa & Ikegaki, 1980; UNEP, 2016), a

health concern that has recently been emphasized to protect human

health by the Minamata Convention on Mercury (UNEP, 2016).

Climate change is also affecting food webs, marine ecosystems,

and the goods and ecological services they provide. The global aver-

age atmospheric carbon dioxide (CO2) concentration has increased

from 278 to >400 parts per million (ppm) since the pre-industrial

era, leading to a range of changes in ocean conditions (IPCC, 2014).

Average ocean temperatures in the upper 75 m have increased at a

rate of over 0.1°C per decade during the period 1971–2010 (IPCC,

2014). Simultaneously, surface ocean acidity increased by approxi-

mately 30% since preindustrial levels (Doney, Fabry, Feely & Kley-

pas, 2009), while oxygen concentration in the open ocean decreased

by 3–5 lmol/kg per decade due to increasing sea surface tempera-

tures and expansion of anaerobic zones along coastal zones (IPCC,

2014). Global warming is projected to reduce oxygen content,

decrease sea ice extent, and increase sea level as the 21st century

progresses (Gattuso et al., 2015). These changes are expected to

alter fish size and distribution (Cheung et al., 2013; Pinksy, Worm,

Fogarty, Sarmiento & Levin, 2013; Poloczanska, Hoegh-Guldberg,

Cheung, P€ortner & Burrows, 2014; Poloczanska et al., 2013), food

web structure (Beaugrand, Edwards, Raybaud, Goberville & Kirby,

2015), trophodynamics (Ainsworth et al., 2011; Kirby & Beaugrand,

2009; Stock, Dunne & John, 2014), and productivity of marine

organisms and ecosystems (Gattuso et al., 2015; P€ortner et al.,

2014). Consequently, fisheries will be impacted through changes in

distribution and potential catches (Barange et al., 2014; Cheung,

Reygondeau & Fr€olicher, 2016) and are likely to impact food security

and economic well-being in the long term (Golden et al., 2016).

While the impacts of climate change and chemical pollutants

have been explored independently, potential interactions in food

webs between these two anthropogenic stressors are not clear (Nik-

inmaa, 2013). Climate change may affect bioaccumulation rates for

contaminants including POPs and MeHg in marine food webs by

altering their transport, persistence, and exposure (e.g., Balbus, Box-

all, Fenske, Mckone & Zeise, 2013; Cossa, 2013; Krabbenhoft &

ALAVA ET AL. | 3985

Sunderland, 2013; Noyes et al., 2009; Schiedek, Sundelin, Readman

& Macdonald, 2007; Stahl et al., 2013; Wenning et al., 2010). For

instance, as arctic sea ice is retreating with global warming (IPCC,

2014), exposure of apex predators to PCBs and mercury may

increase (Bond, Hobson & Branfireun, 2015; Fort et al., 2015;

McKinney, Peacock & Letcher, 2009; McKinney et al., 2012, 2013).

Climate change may therefore increase the level of toxic contami-

nants in species harvested by commercial, recreational, and aborigi-

nal fisheries (e.g., Balbus et al., 2013; Marques, Nunes, Moore &

Strom, 2010; Tirado, Clarke, Jaykus, McQuatters-Gollop & Frank,

2010; Weatherdon, Ota, Jones, Close & Cheung, 2016).

While some evidence points to a climate change-related alter-

ation in contaminant dynamics in marine systems (Balbus et al.,

2013; Grannas et al., 2013; Krabbenhoft & Sunderland, 2013; Mac-

donald, Harner & Fyfe, 2005; Nikinmaa, 2013), little is known about

the consequence of these two interacting stressors on the future of

marine food webs, livelihood of coastal human communities, and

safety of direct or indirect consumption of seafood. We review here

the potential interacting effects of climate change and chemical pol-

lution on marine food webs and implications for coastal aboriginal

communities and the average human consumers; that is, Coastal

Indigenous people consume on average 15 time more seafood per

capita than non-Indigenous people (Cisneros-Montemayor, Pauly,

Weatherdon & Ota, 2016). We then present a conceptual framework

to explore the interplay between these anthropogenic stressors and

the need of ecosystem modeling tools to inform climate change and

pollution management policies. We use case studies to apply the

framework to guide an understanding of PCB and Hg dynamics and

impacts in a changing Northeastern Pacific Ocean.

2 | CLIMATE CHANGE AND POLLUTANTINTERACTIONS IN MARINE FOOD WEBS

We used the Web of Science, Science Direct, and Google Scholar to

search for peer-reviewed articles, using key words to focus our atten-

tion on the interaction of anthropogenic climate change and contami-

nant bioaccumulation. These keywords included “global climate

change”, “ocean warming”, “ocean acidification”, “marine pollution”,

“chemicals”, “contaminants”, “POPs”, “PCBs”, “DDTs,” “mercury”,

“methylmercury”, “marine food webs”, “food web bioaccumulation

models”, “ecosystem models”, “Arctic”, “Antarctic”, “Atlantic”, “Pacific

Ocean”, “tropics”, “wildlife”, “fish”, “marine mammals”, among others.

While this searching effort attempted to be systematic within the

scope of the review focused on the Northeastern Pacific, we included

lines of evidence based on published research from other regions of

the world to take into consideration examples and study cases of the

mechanisms and interactions between climate change and pollutants

in marine food webs. We acknowledge that this may not be an

exhaustive representation of all the material that exists in the

scientific literature on this topic.

Based on the results of our literature review, we found 23

papers were dedicated to the topic of climate–contaminant

bioaccumulation interactions published between 2003 and 2015

(Tables 1 and 2). Nine of the papers (40%) were review articles on

the interplay between climate change and contaminant bioaccumula-

tion, including some field data, in aquatic food webs and impacts on

foraging ecology (Carere, Miniero & Cicero, 2011; Gouin et al.,

2013; Jenssen et al., 2015; Krabbenhoft & Sunderland, 2013; Mac-

donald & Loseto, 2010; Macdonald et al., 2005; McKinney et al.,

2015; Schiedek et al., 2007). The remaining papers (60%) were field

research and aquatic food web bioaccumulation modeling that

focused on Arctic, Antarctic, and temperate regions (Table 2). Six of

these papers were supported with field data, historical and time ser-

ies contaminant data exclusively focused on the Arctic (Braune et al.,

2014, Braune, Gaston, et al., 2015; Jenssen et al., 2015; Macdonald

et al., 2005; McKinney et al., 2015) or based on hypothetical food

web bioaccumulation modeling work (Borg�a, Saloranta & Ruus,

2010; Carere et al., 2011; Gouin et al., 2013).

Interactions between climate change and contaminants are likely

to affect exposure and bioaccumulation for both humans and nonhu-

man biota (Table 2). This may take the form of changes in primary

productivity (Borg�a et al., 2010), potential climate change effects on

pollutant bioaccumulation in marine pelagic food webs (Hallanger,

Ruus, et al., 2011, Hallanger, Warner, et al. 2011), a mismatch in

temperature preferences for predator–prey interactions (Ng & Gray,

2011), and chemical exposure changes in human diets across eco-cli-

matic regions (Undeman, Brown, Wania & McLachlan, 2010). In con-

trast, the direct bioenergetic impacts of temperature on uptake

through consumption rates or on loss processes such as metabolic

elimination or growth dilution are poorly documented (Borg�a et al.,

2010; Carere et al., 2011; Gouin et al., 2013). Conversely, several

other studies have documented effects implying either direct or indi-

rect impacts of both stressors on feeding ecology, including changes

in foraging behavior and dietary preferences with associated

increases in pollutant exposure and bioaccumulation in polar bears

(Ursus maritimus) (McKinney et al., 2009, 2012, 2013, 2015; Jenssen

et al., 2015; Tables 1 and 2), and exacerbated food web bioaccumu-

lation at the individual and population levels (Booth & Zeller, 2005;

Cullon et al., 2009; Geisz, Dickhut, Cochran, Fraser & Ducklow,

2008; Jenssen et al., 2015).

While several empirical studies in biota (marine mammals and

sea birds) from the Arctic have documented increased concentra-

tions of some POPs in polar bears (McKinney et al., 2009, 2012,

2015) and mercury in sea birds (Braune et al., 2014), few field con-

tributions have shown that the concentrations of an important num-

ber of legacy POPs (e.g., PCBs, DDTs, OC pesticides; Table 2) have

actually decreased in food webs of seabirds (Braune, Gaston, et al.,

2015) or exhibited a lack of increase in some polar bear populations

(McKinney et al., 2013). These examples of decreases in contami-

nants concentrations were attributed to shifts in diets and changes

in trophic position, as well as the phasing out of the use of these

contaminants (i.e., PCBs, DDTs) in the late 1970s. In contrast, a con-

comitant increase in recently emerging POPs (PBDE flame retar-

dants) was observed in Greenland polar bears (McKinney et al.,

2013; see Table 2). While some evidence is available for the Arctic,

3986 | ALAVA ET AL.

data for the Northeastern Pacific and other temperate regions are

scarcely available.

As highlighted by Macdonald et al. (2005), climate variables

expressed through prey availability and biological condition (i.e., less

fat reserves) may readily influence POP exposure and accumulation

in apex predators. Similarly, melting glaciers from the western

Antarctic Peninsula are considered as a possible recurrent source of

DDT contamination to the western Antarctic marine food web and

ecosystem, where DDT levels did not change significantly in Adelie

penguins (Pygoscelis adeliae) for more than 30 years (Geisz et al.,

2008). Climate change–contaminant interactions have also been

noted in geographic areas other than polar regions, including the

Great Lakes, the Faroe Islands, the Northeastern Pacific Ocean

(Tables 1 and 2), and the Antarctic continent (Geisz et al., 2008;

TABLE 1 A limited number of research review studies assessed the impact of climate change on the bioaccumulation of chemical pollutantsin marine ecosystems and food webs. Studies are sorted in chronological order

Year Marine food web species/organisms Region Pollutants Findings Reference

2005 Arctic ecosystems, for example,

Calanus hyperboreus, Themisto

libellula, Arctic cod (Arctogadus

glacialis), black guillemot (Cepphus

grylle), glaucous gull (Larus

hyperboreus), ringed seal (Pusa

hispida), polar bear (Ursus maritimus)

Arctic POPs (PCBs, DDTs,

HCHsa), mercury

(MeHg), and

radionuclides

Potential exacerbation of POPs and

mercury in marine food webs due to

climate change (i.e., increasing

temperatures)

Macdonald

et al. (2005)

2007 Cycling of contaminants in aquatic

food webs in general (e.g., Arctic)

Global POPs, PAHsb, and

metals

Climate change impacts can affect

contaminant exposure and

biomagnification in Arctic marine food

webs

Schiedek et al.

(2007)

2010 Seafood systems Global Metals (mercury/

MeHg, cadmium,

lead), POPs, (i.e.,

DDTs) and PAHs

A variety of responses to climate change

across trace metals and organic

contaminants was found in seafood

Marques et al.

(2010)

2010 Arctic food webs Arctic Mercury (MeHg) Climate change can change mercury

cycling and possible regeneration in the

marine environment affecting MeHg

bioaccumulation in food webs

Macdonald

and Loseto

(2010)

2011 Bioaccumulation in aquatic biota in

general

Aquatic

environments

(e.g., marine)

POPs and mercury Climate change can alter contaminants’distribution in water bodies, enhancing

the bioaccumulative pollutants’bioavailability, and increasing risk of

food chain transfer

Carere et al.

(2011)

2013 The round goby (A. melanostomus)

food chain

Hypothetical

multimedia

environment

POP-like chemicals

and hypothetical

chemicals

Food web model predicted small

bioaccumulation responses in an

aquatic food chain, varying

substantially, depending on partitioning

properties and biotransformation rates

Gouin et al.

(2013)

2013 General insights for food webs Global ocean and

environment

Mercury (MeHg) Increases of ocean oxygen minimum

zones and changes in productivity due

to climate change will exacerbate MeHg

production and marine food web

bioaccumulation

Krabbenhoft

and

Sunderland

(2013)

2015 Arctic marine ecosystems/food webs Arctic POPs and mercury Climate change-induced ecological

changes and alterations in POP and

mercury exposures. Lower sea ice

linked-diet changes associated with

higher contamination in polar bears,

ringed seals, and thick-billed murres

McKinney

et al. (2015)

2015 Polar bear (U. maritimus) food web Arctic POPs Potential climate warming and pollutant

exposure interaction impacting polar

bears due to limited access to prey and

sea ice habitat loss, resulting in

prolonged fasting periods and increased

POP levels

Jenssen et al.

(2015)

aHCHs, hexachlorocyclohexanes.bPAHs, polycyclic aromatics hydrocarbons.

ALAVA ET AL. | 3987

TABLE 2 Field studies and modeling work provide insight into the interplay between anthropogenic contaminants and climate change inmarine food webs. Studies are sorted in chronological order

Year/Timeperiod

Marine food web species/organisms Region Pollutants Findings Reference

1975–2013 Arctic food web: Arctic-breeding

seabirds, thick-billed murres (Uria

lomvia)

Canadian Arctic Mercury Total mercury concentrations in

murre eggs increased significantly

at Prince Leopold Island due to

changes in seabird’s diet and

trophic position driven by climate

change

Braune et al.

(2014)

1984–2011 Eastern’s Greenland polar bear

marine food chain: hooded seal

(Cystophora cristata) and harp seal

(Phoca groenlandica)–polar bear

(U. maritimus)

East Greenland,

Arctic

POPs (PCBs, DDTs,

p,p0-DDE, HCB,

Chlordanes

Octachlorostyrene

(OCS), PBDEs,

PBDE 153)

Brominated POPs in polar bears

increased annually due to climate-

induced dietary shift from

nearshore/benthic/ice-associated

seals to more POPs’ contaminated

offshore/open water seals

McKinney et al.

(2013)

1987–1996 Benthic and pelagic food chains:

bottom invertebrate feeder

Gobionotothen gibberifrons; krill

feeder, Champsocephalus gunnari;

fish feeder, Chaenocephalus

aceratus

Antarctic POPs (PCBs, HCB, p,

p0-DDE, mirex,

nonachlor III, trans-

nonachlor,

toxaphene)

Significant increases of POPs in

benthos and fish feeders. Changing

POP levels reflects global

redistribution and transfer to

Antarctic due to climate change

Weber and

Goerke (2003)

1991–2007 Western Hudson Bay’s polar bear

marine food chain: harbor seal

(Phoca vitulina) or/and harp seal

(P. groenlandica)–polar bear

(U. maritimus)

Canadian Arctic POPs (PCBs, PBDEsa

and b-HCHb)

Increased POPs in polar bears due

to sea ice-associated diet changes,

influenced by climate change, by

feeding on more contaminated

open water-associated seal species

McKinney et al.

(2009)

1993–2013 Arctic food web: Arctic-breeding

seabirds, thick-billed murres

(U. lomvia)

Canadian Arctic POPs (PCBs, p,p0-DDE, HCB,

heptachlor epoxide,

oxychlordane,

dieldrin)

Changes in seabird’s diet and

trophic position driven by climate

change affect rates of decline for

organochlorines from 1992 to

2013

Braune, Gaston,

et al., 2015

2000–2001 Regional food chain: Chinook

salmon (Oncorhynchus

tshawytscha)–southern resident

killer whales (Orcinus orca)

Northeastern

Pacific: Strait

of Georgia

and Puget

Sound

POPs (PCBs PCDDsf,

PCDFsg, and DDTs)

Increasing climate-related stresses

on salmon population abundance

with reduced lipid content,

increasing contaminant exposure

for resident killer whales

Cullon et al.

(2009)

2004–2006 Antarctic marine food web: krill

(Euphausia superba and

E. crystallorophias)–Adelie

penguins (Pygoscelis adeliae)

Western

Antarctic

Peninsula

DDT DDT concentrations have not

decreased in Adelie penguins in the

western Antarctic for more than

30 years due to DDT release from

melting glaciers

Geisz et al.

(2008)

2005 Faroe Islands’ marine ecosystem

food web, including the food

chains: cod (Gadus morhua)–

humans (Faroe Islanders); pilot

whale (Globicephala melas)–

humans (Faroe Islanders)

Faroe Islands

(Northeast

Atlantic)

Mercury (MeHg) Under simulations of climate change

scenarios, MeHg increased in the

ecosystem, translating into

increased human exposure due to

whale meat consumption over time

Booth and

Zeller (2005)

2007 Arctic marine pelagic food webs:

zooplankton-fish seabird food

web

European Arctic POPs Future warming of the Arctic and

increased invasion by boreal

species can result in increased food

web magnification of POPs in polar

species, caused by climate change

Hallanger,

Warner et al.

(2011)

2007–2008 Arctic marine food web involving

transient/subarctic species and

resident species

Canadian Arctic POPs (PCB 4, PCB 5,

PCB 6, PCB 7,

trans-nonachlor,

heptachlor epoxide,

p,p0-DDEc, dieldrin,

HCBd).

Climate change, long-range

transport and increase intransient-

subarctic animals with higher POP

levels trigger higher

biomagnification in transient food

webs

McKinney et al.

(2012)

(Continues)

3988 | ALAVA ET AL.

Weber & Goerke, 2003). In the Arctic ecosystems, POPs and mer-

cury were the main pollutants with demonstrated linkages with cli-

mate change (Jenssen et al., 2015). Based on these observations,

PCBs and MeHg predominate as model chemical candidates to track

and predict the bioaccumulation potential under forward-looking cli-

mate change scenarios.

3 | CLIMATE CHANGE–POLLUTANTBIOACCUMULATION INTERACTIONS:TOWARD A CONCEPTUAL FRAMEWORK

Given the limited basis for predicting status and trends for climate–

contaminant interactions into the future, a conceptual framework

across different levels of biological complexity offers a means to

explore future climate–contaminant scenarios (Figure 1). This frame-

work illustrates the impact of ocean warming and acidification on

pollutant bioaccumulation in food webs and fisheries resources. To

achieve this, we identify the major putative mechanisms and interac-

tions in marine biota and food webs, using PCBs and MeHg as

model contaminants.

Because of the temperature and thermodynamic dependence of

the physico-chemical properties of PCBs and mercury, climate

change is expected to impact bioaccumulation in aquatic biota and

biomagnification in aquatic food webs (e.g., Carere et al., 2011;

Gouin et al., 2013; Krabbenhoft & Sunderland, 2013; Macdonald

et al., 2005). The redistribution and increased volatilization of PCBs

and evasion of mercury from the ocean, freshwater, sediments, and

land to the atmosphere due to increasing temperatures are projected

to increase the delivery of these pollutants to the Atlantic, Pacific,

and Arctic regions (e.g., Dalla Valle, Codato & Marcomini, 2007;

Grannas et al., 2013; Krabbenhoft & Sunderland, 2013; Lamon et al.,

2009; Macdonald et al., 2005; MacLeod, Riley & Mckone, 2005;

Strode et al., 2007; Wania, 2003). Recent studies have found that

PCBs deposited in oceans and ice are being remobilized into the

atmosphere over the past two decades as a result of climate change

in the Arctic and Atlantic Ocean (Ma & Hung, 2012; Ma, Hung, Tian

& Kallenborn, 2011; Nizzetto, Lohmann, Gioia, Dachs & Jones,

2010), but overall declines in PCB concentrations are still observed

in the Arctic environment (AMAP, 2014). However, these transport-

partitioning processes and trends can be further aggravated by the

“grasshopper effect” and fractionation during long-range atmospheric

transport for PCBs (Gouin, Mackay, Jones, Harne & Meijer, 2004).

Thus, climate change has the potential to influence pollutant trans-

port (e.g., PCBs and mercury) in the earth system and may alter the

exposure and risks to marine organisms and humans.

TABLE 2 (Continued)

Year/Timeperiod

Marine food web species/organisms Region Pollutants Findings Reference

2008 Arctic and Atlantic fjord systems:

zooplankton species

European Arctic POPs Higher POP concentrations in

zooplankton were due to fjord

specific characteristics (i.e., ice

cover and timing of snow/glacier

melt), possibly caused by climate

change

Hallanger, Ruus,

et al. (2011)

2010 Arctic food web: Copepods

(Calanus glacialis C. hyperboreus;

krill: Thysanoessa inermis); pelagic

amphipods: T. libellula; fish: polar

cod, Boreogadus saida); seabirds:

kittiwake, Rissa tridactyla)

European Arctic PCB 52, PCB 153, a-

HCHe

Food web bioaccumulation model

predicted lower bioaccumulation

under climate change scenario

Borg�a et al.

(2010)

2010 Arctic food web: marine mammals

(seals)-Inuit

Canadian Arctic Hypothetical organic

persistent chemicals,

including POPs

Climate -bioaccumulation

interactions revealed the potential

for increased contaminant

concentrations in Inuit diet

Undeman et al.

(2010)

2011 Great Lakes food web: the round

goby (Apollonia melanostomus);

the mottled sculpin (Cottus

bairdii); the lake trout (Salvelinus

namaycush)

Great Lakes

(Lake Erie and

Lake Superior)

PCB 77 Bioaccumulation projected by food

web modeling was confounded by

predator–prey dynamics due to

temperature preference

mismatches of predator and prey

(growth/uptake rates)

Ng and Gray

(2011)

aPBDEs: Polybrominated diphenyl ethers.bb-HCH: beta-hexachlorocyclohexane.cp,p0-DDE: 4,40-dichlorodiphenyldichloroethylene.dHCB: hexachlorobenzene.ea-HCH: alpha-hexachlorocyclohexane.fPCDDs: polychlorinated dibenzo-p-dioxins.gPCDFs: polychlorinated dibenzofurans.

ALAVA ET AL. | 3989

Reductions in the size or condition of some fish associated with

climate change may increase PCB concentrations in remaining lipid

tissues by counteracting the “growth dilution effect” or biodilution

on chemicals accumulated in fish tissues (e.g., lipid, muscle). The

maximum body weight of fish is limited by dissolved oxygen (DO)

and food resources or energy (Enberg, Dunlop & Jørgensen, 2008).

An increase in temperature and decrease in oxygen content in the

environment is thought to result in fish reaching a smaller maximum

body size (Cheung et al., 2013). Simulation modeling of fish growth

and biogeography projects a decrease in community-level average

body size of fishes of 14–24% by the 2050s relative to the 2000s

period under a “business-as-usual” climate change scenario (Cheung

et al., 2013). Moreover, food consumption rate (and subsequently

the pollutants in the food items) of fish increases as temperature

increases. At a higher temperature, gill ventilation rate increases in

response to an increased metabolic rate and decrease in DO (Ken-

nedy & Walsh, 1997). As a result, increases in consumption and ven-

tilation rate may increase the concentration of pollutants in fish’s

F IGURE 1 Our proposed conceptual framework postulates the impacts of climate change through climate-induced pollutant sensitivity onbioaccumulation in marine food webs. Two major mechanisms (ocean acidification and ocean warming) interact to intensify exposure andpotential bioaccumulation (e.g., produce elevated concentrations in top predators) and enhance effects across hierarchical biological levels oforganization from genes to ecosystems. The ultimate impacts on ecosystem goods and services are captured in this framework, affecting thesocioeconomic and human dimensions. Of particular consideration is also the impact of climate change on contaminant toxicity. For instance,under climate change effects on temperature, salinity, and pH, empirical studies on toxicity testing of pesticides in phytoplankton andinvertebrates have shown increased toxicity by 4–5 times at higher temperatures and salinities relative to standard conditions (see DeLorenzo,Danese & Baird, 2013; DeLorenzo, Wallace, Danese & Baird, 2009). Conversely, pollutant-induced climate change susceptibility conspiresagainst the adaptability of organisms and food webs to make levels of biological organization more vulnerable to climate change risk andimpacts, thus reducing ecosystem resiliency. To assess and predict these impacts, more empirical data and model approaches are needed toproject potential changes and effects in food webs and ecosystems, upon which top predators and humans strongly rely on. The outcome frommodeling work can be used as input data to inform decision-making processes and develop mitigation or adaptation policies for climate changeand chemical pollution. Risk management can be aimed at reducing the impacts of anthropogenic climate change in marine food webs andecosystems. The arrows ↑ and ↓ indicate an increase and a decrease in a given variable or factor, respectively [Colour figure can be viewed atwileyonlinelibrary.com]

3990 | ALAVA ET AL.

body at smaller maximum body size. In addition, if increased pollu-

tant exposure causes increased sublethal metabolic stress, the cost

of maintenance metabolism would be increased further reducing

assimilation efficiencies. In the Northeastern Pacific, Cullon et al.

(2009) suggested that Chinook salmon affected by ocean warming

and with low lipid reserves may be more contaminated by PCBs,

increasing the risk of exposure and bioaccumulation to its major

predator, resident killer whales (Ross, Cullon, Buckman & Ford,

2010).

Finally, climate change and ocean acidification can alter food

web structures by causing changes in primary and secondary pro-

duction, species compositions, predator–prey abundances, and inter-

and intraspecific interactions (e.g., Le Quesne & Pinnegar, 2012;

Nagelkerken & Connell, 2015; Pistevos, Nagelkerken, Rossi, Olmos

& Connell, 2015; Sydeman, Poloczanska, Reed & Thompson, 2015).

As primary production will be affected by climate change, the

amount of POPs such as PCBs readily absorbed by phytoplankton

and amplified through food web will be affected, as well (e.g., Borg�a

et al., 2010; Carere et al., 2011; Gouin et al., 2013; Macdonald,

Mackay & Hickie, 2002; Ng & Gray, 2011; Nikinmaa, 2013; Schie-

dek et al., 2007). The removal or addition of trophic levels and

alteration of bottom-up (i.e., disrupted primary production and

nutrient cycling) or top-down (i.e., reduction or loss of top preda-

tors) mechanisms in the food web mediated by climate change-

induced pollutant sensitivity processes are likely to have dramatic

effects on the bioaccumulation and biomagnification patterns of

PCBs and mercury, that is, increase or decrease in pollutants levels

in the food web (Balbus et al., 2013; Braune, Gaston & Mallory,

2016; Braune et al., 2014, Braune, Gaston, et al., 2015; Gouin

et al., 2013; Jenssen et al., 2015; Krabbenhoft & Sunderland, 2013;

Macdonald et al., 2005; McKinney et al., 2009, 2012, 2013, 2015;

Schiedek et al., 2007).

4 | CASE STUDY: CLIMATE–POLLUTANTIMPACTS ON APEX PREDATORS IN THENORTHEASTERN PACIFIC

We examined studies of PCBs and MeHg in the Northeast Pacific

Ocean to illustrate the potential consequences of climate change–

contaminant interactions on humans and the environment.

4.1 | PCBs as a chemical of concern in climatechange–bioaccumulation interactions

In the Northeastern Pacific, transient and resident killer whales (Orci-

nus orca) continue to be the most PCB-contaminated organisms of

the world, which raises serious concerns about their long-term sur-

vival (Alava et al., 2012; Best et al., 2010; Buckman et al., 2011;

Hickie, Ross, Macdonald & Ford, 2007; Lundin et al., 2016; Ross

et al., 2000). While PCBs represent only one chemical class found in

killer whales, they are considered the preeminent contaminant threat

to high trophic level species in the Northern Hemisphere (Elliott,

Butler, Norstrom & Whitehead, 1989; Ross et al., 2000; Ylitalo et al.,

2001). In addition, ocean temperatures are projected to increase in

the 21st century in the Northeast Pacific, resulting in shifts in distri-

bution and abundance of species, including important prey of killer

whales such as salmon (Cheung, Brodeur, Okey & Pauly, 2015; Cul-

lon et al., 2009; Ford, Ellis, Olesiuk & Balcomb, 2010; Okey, Alidina,

Lo & Jessen, 2014; Ross et al., 2010) (Figure 2). Reduced abundance

of their primary prey, Chinook salmon (Oncorhynchus tshawytscha),

during warming ocean episodes has coincided with subsequent peri-

ods of higher mortality in resident killer whale populations (Ford

et al., 2010). As PCBs are fat-soluble contaminants that amplify in

food webs, productivity-related reduction in lipids may concentrate

PCBs in the diminishing blubber stores of marine mammals (Jenssen

et al., 2015; Lundin et al., 2016; McKinney et al., 2015; Ross et al.,

2010). While pollution risks by PCBs, noise and disturbance, and

reduce prey abundance (i.e., Chinook salmon) are known to affect

these killer whales (Alava et al., 2012; Cullon et al., 2009; Ford et al.,

2010; Hickie et al., 2007; Lundin et al., 2016; Ross et al., 2000), the

combined impact of climate change and PCBs on this top predator

and its prey is poorly understood.

Here, we explore the potential impacts of climate change–pollu-

tant interactions in killer whale food webs through the following

three pathways: (i) climate change-induced pollutant sensitivity; (ii)

pollutant-induced climate change sensitivity; and (iii) a combination

of these two processes.

4.1.1 | Climate change-induced PCB sensitivityscenario

Climate change may increase PCB accumulation in killer whales

through a variety of processes. Survival of killer whales is driven, at

least in part, by the abundance of their key prey species (Ford et al.,

2010). Increased ocean temperatures may reduce lipid content, size,

and abundance of Chinook salmon as a result of reduced primary

productivity and increased physiological stress (e.g., decreased aero-

bic scope, lower growth rate, and high maintenance cost of home-

ostasis/metabolisms) (Crossin et al., 2008; Crozier et al., 2008;

Farrell et al., 2008; Finney, Gregory-Eaves, Sweetman, Douglas &

Smol, 2000; Ford et al., 2010). This may lead to changes in PCB

exposure in resident killer whales (Lundin et al., 2016). For example,

increasing PCB concentrations with diminishing lipid reserves were

observed in Chinook salmon in the Northeastern Pacific (Cullon

et al., 2009; O’Neill & West, 2009; Figure 2). Cullon et al. (2009)

hypothesized that the lower lipid content of southern Chinook

stocks may cause southern resident killer whales to increase their

salmon consumption by as much as 50% potentially increasing expo-

sure to PCBs. This can be referred to as increased climate change-

induced pollutant sensitivity (Hooper et al., 2013; Figure 1). Along

these lines, Rhind (2009) suggested that the interactions between

the effects of pollutants and environmental stressors such as under-

nutrition or osmotic stresses and changes in variables associated

with climatic changes may exacerbate physiological responses to pol-

lutant burdens.

ALAVA ET AL. | 3991

4.1.2 | PCB-induced climate change susceptibilityscenario

PCBs affect the metabolic rate of marine mammals by altering thy-

roid hormone physiology and metabolism (Buckman et al., 2011;

Mos et al., 2010; Tabuchi et al., 2006), potentially increasing their

sensitivity to climate change impacts (Figure 2). Increased PCB con-

centrations in harbor seals and killer whales in the Northeastern

Pacific increase thyroid hormones receptor (i.e., TRa) gene expres-

sion (Buckman et al., 2011; Tabuchi et al., 2006). Subsequently, this

may increase energetic demand, and thus, feeding rate (Ross et al.,

2010). Hooper et al. (2013) highlighted the vulnerability of the thy-

roid system in wildlife to impairment by environmental contaminants,

including PCBs. Consequently, exposures to thyroid-disrupting chem-

icals may impair the ability of vertebrates to adequately respond to

climate changes, that is, a toxicant induced climate sensitivity (Hoo-

per et al., 2013; Figure 1).

Immunotoxicity associated with PCBs (Mos et al., 2010) affects

the ability of animals to combat disease, which is especially relevant

during periods of nutritional stress and other types of stressors in

highly variable environments (e.g., El Ni~no events or El Ni~no South-

ern Oscillation-ENSO) when mass mortality occurs and populations

often approach the critical tipping point of extinction (Alava, Ross,

et al., 2011, Alava, Salazar et al., 2011). Exposure to immunotoxic

contaminants may facilitate the emergence of infectious disease out-

breaks (Desforges et al., 2016; Ross, 2002).

The reproductive impairments and endocrine disruption by PCBs

can have implications at the population level, wherein small threatened

or endangered populations facing bottleneck or Allee effects (i.e.,

decrease in individual fitness at low population size that may result in

critical population thresholds below which populations collapse to

extinction) are most vulnerable, and thus increasing their sensitivity to

climate change. Southern resident killer whales, for example, have a

population size of fewer than 80 individuals and are endangered as a

Figure References:

1. IPCC (2014); 2. MacLeod et al. (2005); 3. Dalla et al. (2007); 4. Lamon et al. (2009); 5. Nizzetto et al. (2010); 6. Ma et al. (2011); 7. Gouin et al. (2013); 8. Grannas et al. (2013); 9. Johannessen et al. (2008); 10. Grant et al. (2011); 11. Cheung et al. (2010); 12. Cheung et al. (2013); 13. Cullon etal. (2009); 14. O'Neill & West (2009); 15. Ross et al. (1996a); 16. Ross et al. (1996b); 17. Tabuchi et al. 2006; 18. Mos et al. (2010); 19. Ross et al. (2010); 20. Buckman et al. (2011); 21. Macdonald et al. (2005); 22. Macdonald & Loseto (2010); 23. Stern et al. (2012); 24. UNEP (2013); 25. Krabbenhoft & Sunderland (2013); 26. Scheuhammer (1991); 27. Siciliano & Lean (2002); 28. Lean (2003); 29. Celo et al. (2006); 30. Merritt & Amirbahman (2009); 31. Riba et al. (2010); 32. Gu et al. (2011); 33. Cossa (2013); 34. De Orte et al. (2014); 35. Simoneau et al. (2005); 36. Reist etal. (2006); 37. Riget et al. (2010); 38. Wiener et al. (2007); 39. Dietz et al. (2013); 40. Scheuhammer et al. (2015); 41. Fort et al. (2015).

F IGURE 2 Pathways and partitioning of PCBs and mercury, including inorganic (HgII) and organic (methylmercury-MeHg) mercury, in the marineenvironment under the influence and impact of climate change. PCBs and mercury enter the coastal marine environment and the ocean from landand atmospheric routes, including anthropogenic urban/industrial wastewater discharges, anthropogenic air emissions from coal-fired plants,agricultural runoff, ocean disposal, and atmospheric transport (i.e., long-range atmospheric transport and global distillation of PCBs andanthropogenic air emissions of gaseous mercury, Hg°). In the ocean, PCBs and mercury can be distributed in abiotic media (i.e., oceanic atmosphere,water, and sediments) and biotic compartments. For the specific case of mercury, HgII partitions in water and sediments, in which it is methylatedto MeHg and accumulated in biotic compartments and biomagnified in marine food webs. A simplified piscivorous–marine mammalian food web isalso illustrated to show the exposure to contamination and bioaccumulation of PCBs and MeHg, which are accumulated and dissolved in blubber/fat tissues (i.e., lipid) and muscle tissues (i.e., proteins), respectively. PCBs and MeHg exposure can be exacerbated by shrinking of prey (fish) andincreased feeding consumption in top predators, driven by climate change. PCBs and MeHg are bioaccumulated mainly through dietary ingestionwith subsequent potential health effects, including immunotoxicity, reproductive impairments, and endocrine disruption (i.e., upregulation of thyroidhormone, triggering metabolic turnover, and food consumption) by PCBs and neurotoxicity, hepatotoxicity, and nephrotoxicity by MeHg

3992 | ALAVA ET AL.

result of several environmental stressors. Understanding the combined

effects of high levels of contaminants and climate change with associ-

ated reductions in prey that may lead to a “tipping point” for this and

other species is critical; such situations would benefit from coherent,

multifaceted management actions to protect these species.

4.1.3 | Combined impact scenario wherein climatechange and PCBs alter contaminant accumulation

Two processes may underlie increased feeding requirements of mar-

ine mammals as described above, including (i) climate-related hypoth-

esis when lipid content in prey may decrease under ocean warming

(i.e., climate change dominant) and (ii) pollutant-related hypothesis

when PCBs may increase metabolic turnover by affecting thyroid hor-

mone physiology (i.e., pollutant dominant). Currently, limited data are

available to identify the likelihood of these hypotheses. A decrease in

body size of prey may also reduce their contaminant concentration as

small fish tend to feed at lower trophic levels relative to larger preda-

tory fish. Even if marine mammals consume more of these prey, the

net accumulation of contaminant may still be lower.

Conversely, both climate change and POPs may interact through

additive or synergetic mechanisms, leading to a need for increased for-

aging effort due to the low abundance of salmon with low lipid

reserves (Cullon et al., 2009; Ford et al., 2010), and/or either

increased ingestion of PCBs because of the upregulation of TRa gene

expression, triggering increasing metabolism and thus increase in

demand for food (Ross et al., 2010; Buckman et al., 2011; Figure 2),

or mobilization of higher PCB concentrations from endogenous lipid

reserves in killer whales during the lowest abundance of prey, that is,

Chinook salmon (Lundin et al., 2016). While prior adaption of some

groups of marine animals or populations to cope with climate change

stress is likely to trigger more resilience and adaptation to such cli-

matic changes (Moe et al., 2013; Parmesan, 2006), there has been lit-

tle time for long-lived species to develop adaptive mechanisms to

cope with industrial PCBs developed in the 20th century (Hooper

et al., 2013; Moe et al., 2013; Stahl et al., 2013). Conversely, for those

species that have been able to adapt to contaminated habitats (Hamil-

ton et al., 2016; Medina, Correa & Barata, 2007), it may be difficult to

confront and adapt to a rapidly changing climate (Hooper et al., 2013;

Moe et al., 2013; Stahl et al., 2013). The interactive effects of a com-

bined climate change-induced pollutant sensitivity and pollutant-

induced climate change susceptibility processes are unclear.

4.2 | Methylmercury as a chemical of concernunder climate change–bioaccumulation interactions

In contrast to PCBs, the climate change-related pollutant sensitivity

and pollutant-related climate change susceptibility processes are less

understood for MeHg. While the cycling and bioaccumulation features

of Hg differ from those governing PCBs, there are some commonali-

ties in the ways in which these two contaminants interact with cli-

mate-related process in the environment and in biota. For instance,

using our salmon killer whale food web from the Northeastern Pacific

(Figure 2), a combination of both climate-induced MeHg sensitivity

and MeHg-induced climate susceptibility processes is possible.

4.2.1 | Impact of climate change on MeHgtransport and fate

The global transport, fate, speciation, and biogeochemical cycling of

mercury can be perturbed by climate change (Macdonald, Mackay, Li

& Hickie, 2003; Macdonald et al., 2005; Macdonald & Loseto, 2010;

Stern et al., 2012; Cossa, 2103; Krabbenhoft & Sunderland, 2013;

Jonsson et al., 2017). The sensitivity of the ecosystem to mercury

may be attributed largely to changes in the methylation processes

within the ice-ocean system rather than depositional changes (Mac-

donald & Loseto, 2010). Once air temperatures rise above the melt-

ing point of ice and snow, concentrations of HgII from atmospheric

deposition (i.e., Hg0 dominates in the atmosphere and is converted

to HgII to be absorbed strongly in particles) are deposited in melting

ice (Krabbenhoft & Sunderland, 2013; Macdonald & Loseto, 2010).

Moreover, blowing/thawing snow and increased terrestrial runoff

can be discharged into the ocean and estuaries, in which it can be

potentially transferred as MeHg in the benthic and pelagic food

webs (Jonsson et al., 2017; Krabbenhoft & Sunderland, 2013; Mac-

donald & Loseto, 2010) or through biogeochemical processes in tem-

perate seas from the North Pacific (Cossa, 2013).

MeHg can be generated in the aquatic environment by biotic and

abiotic pathways, including microbial metabolism and chemical methyla-

tion, respectively (Celo, Lean & Scott, 2006; Schartup et al., 2015). Pro-

duction of MeHg through methylation depends on environmental

factors such as pH, temperature, and the presence of complexing agents

(e.g., chloride). In estuarine and coastal marine environments, microbial

metabolic activity by sulfate-reducing bacteria (SRB) is thought to domi-

nate and control mercury methylation dynamics in anoxic sediments

(Merritt & Amirbahman, 2009), as seen in Figure 2. In addition, it has

recently been shown that zones of nutrient and carbon regeneration in

the ocean (i.e., bacterial regeneration of organic carbon contained in

sinking particles), including the North Pacific, are associated with higher

concentrations of MeHg (Sunderland et al., 2009).These lines of evi-

dence suggest that Hg methylation processes in aquatic environments

may be vulnerable to climate change due to increasing bottom tempera-

tures coupled with more acidic and anoxic sediments. Ultimately, melt-

ing of snow and ice from alpine ecosystems and mountains can release

not only POPs into coastal ecosystems (Morrissey, Bendell-Young &

Elliott, 2005), but potentially mercury, which may enhance exposure to

aquatic organisms in the Northeastern Pacific.

4.2.2 | Impact of climate change on MeHgecotoxicology

Mercury and some metals (e.g., Al, Cu, Fe, Pb, Zn) are generally more

bioavailable in acidified aquatic habitats (Celo et al., 2006; De Orte

et al., 2014; Riba, Kalman, Vale & Blasco, 2010; Roberts et al., 2013;

Scheuhammer, 1991). Acidification not only changes the marine car-

bonate speciation system (i.e., decrease in the concentration of

ALAVA ET AL. | 3993

carbonate ion and the increase in bicarbonate and aqueous CO2),

but alters seawater chemical speciation and biogeochemical cycles of

many elements and compounds (Doney et al., 2009; Roberts et al.,

2013). Mercury concentrations in fish are inversely related to pH in

acidic water (Lean, 2003). Acidification of waters indirectly increases

the exposure of fish-eating carnivores to MeHg, which is likely due

to the higher solubility of mercury in acidic water and enhancement

of mercury methylation rates at lower pH (Lean, 2003). The methyla-

tion process is thought to be mediated by methyltransferase enzyme

pathways in acidic sediments (Siciliano & Lean, 2002). However, this

process can be reduced in anoxic sediments and water by natural

dissolved organic matter (DOM), including humic acid (HA), but only

at low concentrations of these compounds, that is, the reduction in

mercury is inhibited, if HA increases (Gu et al., 2011). Furthermore,

increased global temperatures may increase bacteria metabolic activ-

ity and increased cycling and conversion of mercury from snow melt

into MeHg (Corbitt, Jacob, Holmes, Streets & Sunderland, 2011;

Krabbenhoft & Sunderland, 2013; Macdonald & Loseto, 2010).

The climate change–mercury interaction can be further aggra-

vated by changes in trophic position and the reduction in fish

growth rates, hampering the growth dilution effect, as well as meta-

bolic stress produced by ocean warming and acidification as afore-

mentioned previously, which may sequentially affect or increase

mercury bioaccumulation (Reist et al., 2006; Stern et al., 2012). Fish

with higher growth rates tend to have lower mercury concentrations

(Simoneau, Lucotte, Garceau & Lalibert�e, 2005); thus, fish that are

likely to undergoing shrinking in size and facing homeostasis/meta-

bolism disruption may potentially exhibit increased mercury levels

concentrated in proteinaceous tissues under climate change condi-

tions (Figure 2). In fact, mercury uptake into fish is expected to

increase due to the effect of climate change, that is, increase in tem-

perature (McKinney et al., 2015; Riget, Vorkamp & Muir, 2010).

4.2.3 | Impact of climate change on marine apexpredators and bioaccumulation

The impact of climate change on mercury bioaccumulation in higher

trophic level organisms is best reflected in the pioneer work by Booth

and Zeller (2005). By modeling the trophic transfer and bioaccumula-

tion of MeHg in the Faroe Islands region, these authors predicted

increasing levels of MeHg in this ecosystem under certain climate

change scenarios with an associated increase in human exposure to

this metal from the consumption of pilot whale (Globicephala melas).

Interestingly, the concentrations of mercury have recently increased

in some high trophic level marine biota, including seabirds (e.g.,

Thick-billed murre, Uria lomvia), ringed seals (Pusa hispida), beluga

(Delphinapterus leucas), Greenland shark (Somniosus microcephalus),

and polar bears, in the Canadian Arctic at both temporal and spatial

scales (Braune, Chetelat, et al., 2015). However, while climate change

has been implicated in some increases in mercury concentrations in

Arctic biota, some changes in mercury concentrations observed in

Arctic biota have also included decreases, likely to be influenced by

other nonclimate-related drivers (Braune et al., 2016; McKinney

et al., 2015). Similarly, mesopelagic predators foraging in the North-

eastern Pacific may be at high risk for increasing mercury bioaccumu-

lation, as recently found by Peterson et al. (2015) due to methylation

process in the mesopelagic environment and deep waters.

Conversely, the isotopic composition of fish in the North Pacific

suggests that much of the MeHg entering the marine food web orig-

inates from subsurface waters exhibiting hypoxia (Cossa, 2013). New

research on mercury contamination of the oceanic mesopelagic envi-

ronment and food webs from the Northeastern Pacific has been evi-

denced by the detection of high total mercury (THg) concentrations

(i.e., 5.8–7.0 lg/g) in muscle of deeper-diving and offshore foraging

Northern elephant seals (Mirounga angustirostris) (Peterson et al.,

2015). Similarly, mercury was recently detected in relatively high

concentrations in hair of harbor seals (P. vitulina) from British Colum-

bia and Washington State (USA), with a THg concentrations in the

order of 5.3 � 0.3 lg/g in pups, 4.5 � 0.5 lg/g in juveniles, and

8.3 � 0.8 lg/g in adults, exhibiting the highest THg levels relative to

other marine mammal species from the Northeastern Pacific (No€el

et al., 2015). In muscle tissues of marine mammals and fish, MeHg

generally accounts for an average between 80 and 100% of THg

(Kannan et al., 1998; Magalh~aes et al., 2007; Storelli, Stuffler & Mar-

cotrigiano, 2002; Wagemann, Trebacz, Boila & Lockhart, 1998;

Wagemann, Trebacz, Hunt & Boila, 1997). These recent studies of

Hg, combined with continued atmospheric emissions worldwide

(UNEP, 2013, UNEP, 2016; No€el et al., 2015), highlight continued

concerns about this toxic element. This concern is particularly critical

for coastal aboriginal communities that rely heavily on marine foods

as an important source of nutrition.

5 | MODELS AS TOOLS TO ASSESSCLIMATE–POLLUTANTS INTERACTIONS

Assessing the interactions between climate change and contaminants

is a challenging task. While field observations and empirical work in

laboratory settings using lower trophic level organisms can provide

empirical data to understand the effects of these anthropogenic dri-

vers, research is costly and often constrained by ethical and legal

restrictions when assessing effects in many top level predators (e.g.,

threatened or endangered species). This is where models can serve

as useful tools to predict the impacts of anthropogenic stressors

under different scenarios. Model outputs can be used to generate

hypotheses that field and laboratory studies could focus on, thus

increasing the efficiency of research resources as well as relevancy

to address policy needs. Finally, models can help identify data gaps

and research priorities.

The development of food web bioaccumulation models (Alava,

Ross & Gobas, 2016; Alava et al., 2012; Gobas & Arnot, 2010), ecosys-

tem-based models (e.g., Ecopath) (Ainsworth et al., 2011; Gu�enette,

Ara�ujo & Bundy, 2014), and bioclimatic models (Ainsworth et al.,

2011; Cheung, Dunne, Sarmiento & Pauly, 2011; Cheung et al., 2009,

2010, 2013) can help address aspects of this approach. A combination

of food web bioaccumulation models and trophodynamic models

3994 | ALAVA ET AL.

(Christensen & Pauly, 1992; Christensen & Walter, 2004; Coll�eter

et al., 2015; Pauly, Christensen & Carl Walters, 2000) may offer

unique means to explore climate change–pollutant scenarios (Fig-

ure 2). For example, in Ecopath with Ecosim (EwE), a trophodynamic

modeling approach that is widely applied to study marine ecosystems

in the world, it is possible to link a module to examine bioaccumulation

of contaminants called Ecotracer (Christensen & Walter, 2004; Col-

l�eter et al., 2015; Pauly et al., 2000). The EWE software is user-friendly,

free, and downloadable online (www.ecopath.org).

Ecotracer predicts movement and accumulation of contaminants

in food webs (Booth & Zeller, 2005; Christensen & Walter, 2004;

Christensen, Walters, Pauly & Forrest, 2008; Coll�eter et al., 2015;

Coombs, 2004). Specifically, changes in concentrations of chemicals

(e.g., POPs) are predicted using information about flows of biomass

and contaminants across the food web (e.g., isotope decay rate and

physical exchange rates) (Christensen et al., 2008). Ecotracer has

recently been used to track, simulate, and assess the transfer and

bioaccumulation of POPs such as PCBs in the marine ecosystems of

the eastern Bering Sea (Coombs, 2004), and other pollutants, includ-

ing MeHg exposure and flow in the Faroe Islands marine ecosystems

under present conditions and climate change scenarios (Booth & Zel-

ler, 2005). In addition, a global model to illustrate the movement of

dioxins in the marine food web of the global ocean, using the spatial

trophodynamic model (Ecospace) and Ecotracer, was developed by

Christensen and Booth (Christensen & Booth, 2006). Changes in pri-

mary productivity, species range shifts, zooplankton community size

structure, ocean acidification, and ocean deoxygenation have

recently been assessed using five Ecopath with Ecosim models (Ains-

worth et al., 2011; Gu�enette et al., 2014). These models can help to

elucidate the following research areas:

• Ecological questions regarding the cumulative impact of climate

change and ocean pollution in marine food webs;

• The role and impact of increasing climate change and contami-

nants on marine food web bioaccumulation and associated

effects;

• The effects of contaminants on aquatic biota and on coastal com-

munities under forward-looking scenarios of climate change and

pollutants;

• Exploration of management policy options to address the impact

of climate change and pollution in marine wildlife and coastal

communities; and

• Assessment of economic impacts associated with the food web

and ecosystem effect by climate change and ocean pollution in

terms of food security and benefits of fisheries.

While modeling approaches are useful to generate and test

hypotheses of interactions between climate change and pollutants

(Calosi, De Wit, Thor & Dupont, 2016), empirical data are still of

paramount importance to measure and track the ecological changes

of predator–prey interactions and their impacts on PCB and mercury

concentrations in fish and other wildlife populations (Letcher et al.,

2010; McKinney et al., 2015). On the other hand, outputs from

modeling analysis can inform experimental design so that data can

be collected effectively to examine key aspects of biological systems

responses to the global change (Calosi et al., 2016).

6 | DISCUSSION

Several studies (Tables 1 and 2) suggest that climate change is

interacting with contaminants to intensify or lessen food web

bioaccumulation for PCBs and Hg. This has serious implications at

the top of food webs. There is evidence that climate change may

lead to change contaminant burdens in fish and marine mammals,

with a concomitant decrease in fish quality and nutritional value for

human consumption. While species inhabiting industrialized areas

may be most at risk for contaminants, the particular vulnerability of

the Arctic climate suggests that the latter region may be particularly

vulnerable to climate–contaminant interactions. Marine mammal spe-

cies and wildlife populations from the Arctic exceeded an overall

threshold level of concern (i.e., biological effects in relation to POP

exposure to evaluate risks) of 1 ppm (mg/kg), but a certain evidence

of a POP-associated stress in these populations has yet to be cor-

roborated (Letcher et al., 2010). The social, cultural, nutritional, and

economic benefits of Inuit country foods or traditional foods remain

a topic of significant concern (Kirk et al., 2012; Van Oostdam et al.,

2005).

Climate change is likely to alter the degree of human exposure

to pollutants and the response of human populations to these expo-

sures (Balbus et al., 2013). Changing human behavior will also affect

how humans come into contact with contaminated air, water, and

food in an era of climate change. Climate change and variability may

have an impact on the occurrence of food safety hazards, including

changes in transport pathways for contaminants, at various stages of

the food chain, from primary production through to consumption

(Marques et al., 2010; Tirado et al., 2010). As shown in our frame-

work (Figure 1), these changes may also affect socioeconomic

aspects related to food systems such as agriculture, animal produc-

tion, global trade, demographics, and human behavior which all influ-

ence food safety (Tirado et al., 2010). This has important

implications for human populations in some vulnerable regions such

as the Arctic where exposure to potentially harmful contaminants

such as mercury (i.e., MeHg) and POPs, mainly through traditional

seafoods and a diet rich in marine mammals, is particularly high (Kirk

et al., 2012; Riget et al., 2010; Van Oostdam et al., 2005).

The world fisheries economy is not only threatened by overfish-

ing and habitat degradation, but by pollutants and climate change

(Sumaila, Cheung, Lam, Pauly & Herrick, 2011). For instance, global

fisheries can expect $10-billion revenue loss due to climate change

under a worst-case scenario in the future (Lam, Cheung, Reygondeau

& Sumaila, 2016). Likewise, marine fisheries that depend on small

and large pelagic fish may face socioeconomic challenges due to

degradation of fish product quality that have been contaminated by

pollutants (e.g., mercury). If fish metabolism and quality are changed

due to climate change–contaminant interactions, this could affect

ALAVA ET AL. | 3995

their nutritional value (e.g., omega 3 fatty acids) important in human

diets. In some cases, this may raise serious food safety concerns and

affect consumer demand. Simultaneously, the global ocean supplies a

myriad of vital services that are of global significance, but coming

with an increasing marginal cost caused by anthropogenic activities,

which should be considered when contemplating adaptation to and

mitigation of anthropogenic climate change (Stocker, 2015). Thus,

the protection of marine food webs from climate change and con-

tamination is critical to the conservation of marine species.

Climate change and pollutants have the potential to dramatically

shape the nature of the world’s oceans, but each can be expected

to influence the other in a variety of ways. While evidence on cli-

mate change–pollutant interactions and their impacts on ecosystems

and fisheries are limited, the development of policy actions based

on the best available scientific and local knowledge in close conjunc-

tion with the precautionary principle is urgently required. Also, the

implications of climate change–pollution interactions for international

policy and conventions that regulate marine contaminants should be

evaluated to identify potential gaps and actions. This include, for

example, the Minamata Convention on Mercury that provides the

international legal basis for banning new mercury mines, phasing

out of existing ones, controlling air emissions, and regulating the

informal sectors of artisanal and small-scale gold mining (UNEP,

2016).

This review underscores the potential for an underestimation of

the impacts of climate change and pollutants on marine ecosystems,

because interactions have generally not been considered. However,

this study identifies the various pathways wherein these two stres-

sors may interact. This highlights the need to better understand such

interactions, in support of changes to current mitigation and man-

agement policies and actions. In summary, we (i) identify PCBs and

methylmercury as priority contaminants to consider in future climate

change–pollutant scenarios; (ii) provide a conceptual framework to

explore future scenarios, mitigation plans or adaptation options; and

(iii) propose that an integrated approach to human and ecological

assessments be carried out, in a manner that protects marine wildlife

and coastal aboriginal communities that rely heavily on marine

resources.

ACKNOWLEDGEMENTS

Funding for J.J. Alava was provided by the Mitacs-SSHRC joint ini-

tiative, OceanCanada Partnership and Coastal Ocean Research Insti-

tute (Vancouver Aquarium Marine Science Centre). W.W.L.C.

acknowledges funding support from the Nippon Foundation-Nereus

Program and Natural Sciences and Engineering Research Council of

Canada (NSERC). Special thanks to the Changing Ocean and Fish-

eries Economic Research Units (Institute for the Oceans and Fish-

eries, University of British Columbia) for further academic support.

Special thanks to Dr. G.I. Scott for his valuable insights and com-

ments. The authors declare that there is no conflict of interests

regarding this review manuscript.

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