Please submit your detailed outline (2-3 pages + full-text PDF of each of the 10 articles in the references) here follow these instructions:

The outline must have all major sections of the paper and sub topics for each section. See the following link. How to prepare detailed outline Actions
Each statement needs to have a citation. You must include and cite at least 10 peer-reviewed articles, 7 of which must be primary literature. References must be properly formatted: Citing References in Scientific Papers x Actions Recommended: use reference manager software to handle your citations. There are free services such as Mendeley (Links to an external site.).
Please use this Sample outline Actions to help guide you on what your outline should look like.

Must include 10 full-text PDF of the cited peer-review articles.

Topic The topic I’ve chosen is the “climate change coral reefs algae”. I’m going to address the capacity of coral macroalgal difficulties to recover and the importance to marine life. The effect of temperature, carbon dioxide and benthic algae has on the algo community ecological process. And the negative and positive ways humans are contributing to this fact.

REVIEW

OCEAN CLIMATE CHANGE, PHYTOPLANKTON COMMUNITY RESPONSES, AND
HARMFUL ALGAL BLOOMS: A FORMIDABLE PREDICTIVE CHALLENGE1

Gustaaf M. Hallegraeff 2

Institute of Marine and Antarctic Studies, and School of Plant Science, University of Tasmania, Private Bag 55, Hobart,

Tasmania 7001, Australia

Prediction of the impact of global climate chang

e

on marine HABs is fraught with difficulties. How-
ever, we can learn important lessons from the fossil
record of dinoflagellate cysts; long-term monitoring
programs, such as the Continuous Plankton Recor-
der surveys; and short-term phytoplankton commu-
nity responses to El Niño Southern Oscillatio

n

(ENSO) and North Atlantic Oscillation (NAO) epi-
sodes. Increasing temperature, enhanced surface
stratification, alteration of ocean currents, intensifi-
cation or weakening of local nutrient upwelling,
stimulation of photosynthesis by elevated CO2,
reduced calcification through ocean acidification
(‘‘the other CO2 problem’’), and heavy precipitation
and storm events causing changes in land runoff
and micronutrient availability may all produce con-
tradictory species- or even strain-specific responses.
Complex factor interactions exist, and simulated
ecophysiological laboratory experiments rarely allow
for sufficient acclimation and rarely take into
account physiological plasticity and genetic strain
diversity. We can expect: (i) range expansion of
warm-water species at the expense of cold-water spe-
cies, which are driven poleward; (ii) species-
specific changes in the abundance and seasonal
window of growth of HAB taxa; (iii) earlier timing of
peak production of some phytoplankton; and (iv)
secondary effects for marine food webs, notably
when individual zooplankton and fish grazers are dif-
ferentially impacted (‘‘match-mismatch’’) by clima

te

change. Some species of harmful algae (e.g., toxic
dinoflagellates benefitting from land runoff and ⁄ or
water column stratification, tropical benthic dinofla-
gellates responding to increased water temperature

s

and coral reef disturbance) may become more suc-
cessful, while others may diminish in areas currently
impacted. Our limited understanding of marine eco-
system responses to multifactorial physicochemical
climate drivers as well as our poor knowledge of the
potential of marine microalgae to adapt genetically
and phenotypically to the unprecedented pace of
current climate change are emphasized. The greatest

problems for human society will be caused by being
unprepared for significant range expansions or the
increase of algal biotoxin problems in currently
poorly monitored areas, thus calling for increased
vigilance in seafood-biotoxin and HAB monitoring
programs. Changes in phytoplankton communities
provide a sensitive early warning for climate-driven
perturbations to marine ecosystems.

Key index words: adaptation; algal blooms; climate
change; continuous plankton recorder; ENSO;
NAO; ocean acidification; range expansion

Abbreviations: DMS, dimethylsulfoxide; ENSO, El
Niño-Southern Oscillation; GOOS, Global Ocean
Observation Systems; IMOS, Integrated Marine
Observing System; IPCC, Intergovernmental
Panel on Climate Change; NAO, North Atlantic
Oscillation episodes; OOI, Ocean Observatories
Initiative

HABs in a strict sense are completely natural phe-
nomena that have occurred throughout recorded
history. Even nontoxic algal blooms can have devas-
tating impacts, for instance, when they lead to kills
of fish and invertebrates by generating anoxic con-
ditions in sheltered bays. Other algal species, even
though nontoxic to humans, can produce exudates
or reactive oxygen species that can damage the gill
tissues of fish (raphidophytes Chattonella and Hetero-
sigma, and dinoflagellates Cochlodinium, Karenia, and
Karlodinium). Whereas wild fish stocks can swim
away from problem areas, caged fish in intensive
aquaculture operations are trapped and thus can
suffer catastrophic mortalities. Of greatest concern
to human society are algal species that produce
potent neurotoxins that can find their way through
shellfish and fish to human consumers where they
produce a variety of gastrointestinal and neurologi-
cal illnesses. One of the first recorded fatal cases of
food poisoning after eating contaminated shellfish
happened in 1793, when English surveyor Captain
George Vancouver and his crew landed in British
Columbia (Canada) in an area now known as
Poison Cove. He noted that, for local Indian tribes,

1Received 29 March 2009. Accepted 10 September 2009.
2Author for correspondence: e-mail hallegraeff@utas.edu.au.

J. Phycol. 46,

220

–235 (2010)
� 2010 Phycological Society of America
DOI: 10.1111/j.1529-8817.2010.00815.x

220

it was taboo to eat shellfish when the seawater
became bioluminescent due to algal blooms by the
local dinoflagellate Alexandrium catenella, which we
now know to be a producer of paralytic shellfish
poisons (PSP) (Dale and Yentsch 1978).

The increase in shellfish farming worldwide is
leading to more reports of paralytic, diarrhetic (first
documented in 1976 in Japan), neurotoxic
(reported from the Gulf of Mexico as early as
1840), amnesic (first identified in 1987 in Canada),
or azaspiracid shellfish poising (first identified in
1998 in Ireland). The English explorer Captain
James Cook already suffered from the tropical ill-
ness of ciguatera fish poisoning when he visited
New Caledonia in 1774. Worldwide, close to 2,000
cases of food poisoning from consumption of con-
taminated fish or shellfish are reported each year.
Some 15% of these cases will prove fatal. If not con-
trolled, the economic damage through the slump in
local consumption and export of seafood products
can be considerable. Whales and porpoises can also
become victims when they take up toxins through
the food chain via contaminated zooplankton or
fish. In the USA, poisoning of manatees in Florida
via seagrasses and their faunal epiphytes, and, in
California, of pelicans and sea lions via contami-
nated anchovies have also been reported. In the
past three decades, HABs seem to have become
more frequent, more intense, and more widespread
(Hallegraeff 1993, Van Dolah 2000). There is no
doubt that the growing interest in using coastal
waters for aquaculture is leading to a greater aware-
ness of toxic algal species. People responsible for
deciding quotas for pollutant loadings of coastal
waters, or for managing agriculture and deforesta-
tion, should be made aware that one probable out-
come of allowing polluting chemicals to seep into
the environment will be an increase in HABs or a
change in community structure affecting a change
in food web. In countries that pride themselves on
having disease- and pollution-free aquaculture, every
effort should be made to quarantine sensitive aqua-
culture areas against the unintentional introduction
of nonindigenous harmful algal species. Nor can
any aquaculture industry afford not to monitor for
an increasing number of harmful algal species in
water samples and for an increasing number of algal
toxins in seafood products using increasingly sophis-
ticated analytical techniques. Last but not least, glo-
bal climate change is now adding a new level of
uncertainty to many seafood safety and HAB moni-
toring programs. Whereas in the past two decades
unexpected new algal bloom phenomena have often
been attributed to eutrophication (Smayda 1990) or
ballast water introduction (Hallegraeff 1993, Lilly
et al. 2002), increasingly novel algal bloom episodes
are now circumstantially linked to climate change.

The Intergovernmental Panel on Climate Change
(IPCC 2008) is planning to include HAB risk fore-
casts under a range of climate change scenarios. A

number of scattered publications have started to
address the topic of HABs and climate change, but
they usually have focused on single environmental
factors (e.g., CO2, temperature increase, stratifica-
tion), single biological properties (photosynthesis,
Beardall and Stojkovic 2006; calcification, Rost and
Riebesell 2004; nutrient uptake, Falkowski and Oli-
ver 2007), or addressed selected species categories
of regional interest only (Peperzak 2005, Moore
et al. 2008b). Complex factor interactions are rarely
considered in climate simulation scenarios, and eco-
physiological experiments rarely cover the full range
of genetic diversity and physiological plasticity of
microalgal taxa. Prediction of the impact of global
climate change on algal blooms is fraught with
uncertainties. It is unfortunate that so few long-term
records exist of algal blooms at any single locality,
where ideally we need at least 30 consecutive years.
However, we can learn important lessons from the
dinoflagellate cyst fossil record (Dale 2001), from
the few long-term data sets available, such as the
Continuous Plankton Recorder surveys (Hays et al.
2005) and short-term phytoplankton community
responses to ENSO and NAO episodes. Whereas the
Continuous Plankton Recorder surveys were initially
designed to primarily sample zooplankton, these
instruments also collect phytoplankton down to
even coccolithophorids (Hays et al. 1995). Started
in 1931 in the North Atlantic, gradually these sur-
veys have expanded to the North Pacific (since
1997) and, more recently, also the Western Atlantic,
Australia, and the Southern Ocean (Richardson
et al. 2006). The present review seeks to provide a
broad overview of the complexity of climate variabil-
ity and factor interactions, examine marine phyto-
plankton responses with a focus on the HAB species
niche, and identify major research gaps.

The global climate system. The term ‘‘climate’’ is
used here to include both anthropogenic climate
change as well as the large-scale decadal oceano-
graphic patterns such as the ENSO, Pacific Decadal
Oscillation (PDO), and NAO. This use is in contrast
to ‘‘weather,’’ which occurs over short timescales of
days to weeks (cf. Moore et al. 2008b). The earth’s
climate system comprises the atmosphere (air, water
vapor, constituent gases, clouds, particles), hydro-
sphere (oceans, lakes, rivers, groundwater), and cry-
osphere (continental ice sheets, mountain glaciers,
sea ice, surface snow cover). The oceans are a core
component of the global climate system because
they store 93% (=39,100 gigatonnes, Gt) of the
world’s carbon, but more and more, we are now
becoming aware of the quantitative contribution to
climate by marine phytoplankton, accounting for
50% of global primary productivity (Longhurst et al.
1995). All the microalgal cells in the world oceans
could be packed in a plank, 386,000 km long, 7 cm
thick, and 30 cm wide, that is, stretching from the
earth to the moon (Andersen 2005). This increased
recognition of phytoplankton as a climate driver is

C L I M A T E C H A N G E A N D A L G A L B L O O M S 221

well demonstrated by the commercial interests in
ocean fertilization to combat anthropogenic climate
change (Glibert et al. 2008). Annually, the oceans
absorb 1.8 Gt of carbon through photosynthesis and
2 Gt via abiotic absorption. The oceans thus have
acted as a sink for 30% of all anthropogenic carbon
emissions since the onset of the Industrial Revolu-
tion. The oceans are particularly effective in absorb-
ing heat and have taken up >90% of the increase in
heat content of the earth since 1961. Climate
change in the past has occurred naturally due to
internal fluctuations in the atmosphere, hydro-
sphere, and cryosphere, but it has also been influ-
enced by volcanic eruptions, variations in the sun’s
output, the earth’s orbital variation, and change in
the solid earth (e.g., continental drift).

Climate on our planet has been constantly chang-
ing, over scales of both millions of years (glacial to
interglacial periods) and short-term oscillations of
tens of years (ENSO, NAO). The earth’s climate in
the distant past has at times been subject to much
higher ultraviolet-B (UVB) levels and CO2 concen-
trations than we are seeing at present. The first pho-
tosynthetic cyanobacteria evolved 3.5 billion years
ago at CO2 levels 1,000· those of the present, fol-
lowed by green algae 1,000 million years ago (mya;
500· present) and dinoflagellates 330–400 mya (8·
present), whereas more recently evolved diatoms
and haptophytes operated under comparatively low
CO2 environments (2–3· present) (Beardall and
Raven 2004; Fig. 1).

During the past 800,000 years, atmospheric CO2
has fluctuated between 180 ppm in glacial and
300 ppm in interglacial periods, but in the past
200 years, this has increased from 280 ppm to
>380 ppm at present, with values of 750–1,000 ppm
predicted by 2100. In the past 1,000 years, our
planet has gone through episodes warmer than
present, such as the medieval warm period AD 550–
1300, and colder than now, such as the little ice age
AD 1300–1900. Global temperatures in the past 20–
30 years (Fig. 1, bottom) have increased significantly
with a further rise of 2�C–4(6)�C predicted over the
next 100 years. Undoubtedly, climate change of the
magnitude that we will be experiencing in the next
100 years has happened before, albeit in the past
proceeding at a much slower pace and starting from
a cooler baseline than present (IPCC 2008). Past
episodes of climate change over long periods of
geological and evolutionary history allowed organ-
isms to adapt to their changing environment.
Because of their short generation times and longev-
ity, many phytoplankton are expected to respond to
current climate change with only a very small time
lag. They are expected to spread quickly with mov-
ing water masses into climatic conditions that match
the temperature, salinity, land runoff, and turbu-
lence requirements of the species. However, our
knowledge of the potential of marine microalgae to
adapt is very limited. Collins and Bell (2004) grew

the freshwater microscopic alga Chlamydomonas over
1,000 generations at almost 3· present atmospheric
CO2 concentration. The cells acclimated to the
change but did not show any genetic mutations that
could be described as adaptation.

Defining the niche of HABs. Most HABs are more
or less monospecific events, and the autecology of
the causative organisms thus becomes crucial in
understanding the factors that trigger these phe-
nomena. Defining the niche of key HAB species is
crucial when trying to predict winners or losers
from climate change. An implicit assumption in
ecological studies is that there exists a critical

Fig. 1. Climate change is a matter of scale and time and
can be viewed in terms of thousands of millions of years [evolu-
tion of life on our planet (top), after Beardall and Raven 2004],
hundreds of thousands of years [glacial-interglacial periods, from
Vostok Ice Core data (middle), Lorius et al. 1990], or the past
hundred years [from Hadley Centre for Climate Prediction and
Research, (bottom)]. From a geological perspective, there is
nothing remarkable about the magnitude of climate change we
are experiencing now, except that it appears to proceed at a fas-
ter pace and starts from a warmer baseline.

222 G U S T A A F M . H A L L E G R A E F F

relationship between form and function in organ-
isms, and that life-form therefore is a better predic-
tor of fitness than phylogenetic affinities. Overall,
morphotaxonomy has worked well with HAB spe-
cies, but it is increasingly obvious that ecophysiologi-
cal experiments based on single culture strains can
be highly misleading (Burkholder and Glibert
2006). The development in the past three decades
of the discipline of HAB ecology is evidenced by the
increased frequency and size of international meet-
ings since the first HAB meeting in 1974 and the
creation in 2002 of the dedicated journal Harmful
Algae. A first major review of harmful algal ecology
was produced as a result of a Bermuda NATO-ASI
workshop in 1996 (Anderson et al. 1998), followed
by an update by Graneli and Turner (2006). The
brief summary below (see also Table 1) is largely
based on these two sources.

The commonality of the PSP-producing dinofla-
gellates Alexandrium, Pyrodinium, and Gymnodinium
catenatum lies in the absence of a rapid growth strat-
egy and reliance on benthic resting cysts in life-
cycle transitions (Hallegraeff 1998). Alexandrium
does not usually produce dense biomass blooms
that persist throughout the year. Instead, seasonal
bloom events appear to be restricted in time by cyst
production (Anderson 1997). The persistence of
these cysts through long-term unfavorable condi-
tions allows these dinoflagellates to colonize a wide
spectrum of habitats and hydrographic regimes.
The tropical dinoflagellate Pyrodinium prefers high
salinities (30&–35&) and high temperatures
(25�C–28�C) (Azanza and Taylor 2001). A soil
extract requirement in culture may explain this spe-
cies’ association with rainfall events and land runoff
from mangrove areas. Benthic cyst stages of
G. catenatum (short dormancy period of 2 weeks) do
not play a role in seasonal bloom dynamics, and
their major function is to sustain this species
through long periods when water column condi-
tions are unfavorable for bloom formation (Halle-
graeff et al. 1995). The success of the haptophyte
Phaeocystis in marine systems has been attributed to
its ability to form large gelatinous colonies during
its life cycle. These colonies occupy the same niche
in turbulent, tidally or seasonally mixed water col-
umns as colony-forming spring diatom blooms
(Smayda and Reynolds 2001). The fish killers Hetero-
sigma, Chattonella, Prymnesium, Chrysochromulina, and
Karenia mikimotoi have in common the production
of high biomass blooms together with the produc-
tion of allelopathic chemicals (including reactive
oxygen species) that play a role in predator avoid-
ance (Hallegraeff 1998). Raphidophyte blooms of
Heterosigma are sensitive to temperature for cyst ger-
mination, but chemical conditioning of the water
by land runoff and other growth promoters (e.g.,
from aquaculture wastes) determines the outcome
of competition with diatoms. Similarly, the raphido-
phyte Chattonella includes a benthic cyst stage in its

life history, but the growth of the germling cells as
affected by nutrient conditions and the presence of
diatom competitors holds the key to its bloom
development. The capacity of Chattonella to undergo
vertical migration in stratified water columns with a
shallow nutricline (i.e., nutrients available only at
depth under dim light) provides a competitive
advantage (Imai et al. 1998). While harmful marine
blooms of Chrysochromulina appear to be exceptional
events (in Scandinavia in 1988 and 1991), fish-kill-
ing Prymnesium bloom events in inshore (low salin-
ity), eutrophic waters are recurrent in many parts of
the world. The expression of toxicity by Chrysochrom-
ulina and Prymnesium is variable and can be
enhanced by phosphate limitation (Graneli and
Turner 2006). The fish-killing dinoflagellate Karenia
brevis is a K-strategist, adapted to low nutrient, oligo-
trophic environments. Blooms in the Gulf of Mex-
ico are initiated offshore before being transported
into nearshore waters where they cause fish kills,
discolored water, human respiratory irritation, and
occasionally neurotoxic shellfish poisoning (NSP)
in human shellfish consumers. Taxonomically
related dinoflagellate species of the eurythermal
and euryhaline K. mikimotoi species complex are
associated with marine fauna kills but not human
intoxications. Poorly characterized lipophilic exo-
toxins and mucus production play an allelopathic
role against other algae and also act as agents that
repress zooplankton grazing (Gentien 1998). This
species is especially successful in frontal regions and
in stratified water columns where it accumulates in
the pycnocline (often also a nutricline), thriving on
regenerated ammonia and benefitting from poly-
amine growth factors from decaying diatoms.
Recent success in culturing the dinoflagellate
Dinophysis has confirmed its mixotrophic feeding
behavior on cryptomonad and ⁄ or Mesodinium prey
(Park et al. 2006) and pointed out that the inci-
dence of occasionally high biomass is the result of
active growth and not passive cell accumulation.
The unusual, large phagotrophic dinoflagellate Noc-
tiluca depends upon high prey biomass (mostly dia-
toms) and optimal water temperatures during the
prebloom stage, with starved cells coming to the
surface and aggregating at fronts during calm
weather conditions and wind mixing terminating
the blooms. Diatom blooms of the cosmopolitan
genus Pseudo-nitzschia are common in coastal waters
all over the world. Blooms generally occur during
colder seasons, and seed populations can derive
from both inshore or offshore waters (Bates et al.
1998). The community dynamics of epiphytic ⁄
benthic tropical Gambierdiscus ciguatera dinoflagel-
lates and their associated macroalgal canopy are dic-
tated to a large extent by the degree of water
movement, with other physical and chemical factors
such as temperature, salinity, gases, and inorganic
and organic nutrients only playing a role with
diminishing hydrodynamics (Bagnis et al. 1985).

C L I M A T E C H A N G E A N D A L G A L B L O O M S 223

T
a

b
l
e

1
.

S
u

m
m

a
ry

o
f

p
h

ys
ic

o
ch

e
m

ic

a
l

a
n

d
b

io
lo

g
ic

a
l

n
ic

h
e
-d

e
fi

n
in

g
fa

ct
o

rs
fo

r
k

e
y

H
A

B
sp

e
ci

e
s.

K
e
y

H
A
B
sp
e
ci

e
s

a
n

d
th

e
ir

i

m
p

a
ct

T
e
m

p

e
ra

tu
re

S
a
li

n
it

y
In

o
rg

a
n

ic
n

u
tr

i

e
n

ts
M

ic
ro

n
u

tr
ie

n
ts

(l
a
n

d
ru

n
o

ff
)

O
rg

a
n
ic
n
u
tr

it
io

n
(m

ix
o

t

r
o

p
h

y)
A

ll
e
lo

p
a
th

y;
g

ra
ze

r
a
vo

id
a
n

ce
L

if
e

h
is

to
ry

S
e

n

si
ti

ve
to

tu
rb

u
le

n
ce

A
le

xa
n

d
ri

u
m

(

P
S

P
)

C
ys

t
g

e
rm

i

n
a

ti

o
n

;
b

lo
o

m
w

i

n
d

o
w

F

e
,

C
u

+

N

o
ta

b
ly

ch
a
in

fo
rm

e
rs

G
ym

n
od

in
iu

m
ca

te
n

a
tu

m
(P

S
P

)
B

lo
o
m
w
in
d
o
w
S
e

+
+

P
yr

od
in

iu
m

ba
h

a
m

en
se

(P
S
P
)

T
ro

p
ic

a
l

S
e
;

m
a
n

g
ro

ve
ru

n
o

ff
+

+

C
h

a
tt

on
el

l

a
(fi

sh
k

il
ls

)
C

ys
t

g
e

rm

in
a
ti
o
n

M
n

,
F

e
,
C
u

R

e
sp

o
n

d
s

to
a
q

u
a
cu

lt
u

re
w

a
st

e
s
+

H
et

er
os

ig
m

a
(fi
sh
k
il
ls
)
C
ys
t
g
e
rm
in
a
ti
o
n

F
re

s

h
w

a
te

r
o

n
to

p
o

f
w

a
te

r
co

lu
m

n
M
n
,
F
e
,
C
u
R
e
sp
o
n
d
s
to
a
q
u
a
cu
lt
u
re
w
a
st
e
s
+
+

C
ya

n
o

b
a
ct

e
ri

a
B

lo
o
m
w
in
d
o
w

P
h

o
sp

h
a
te

+
M

ic
ro

cy
st

is
(m

ic
ro
cy
st

in
s)

F
re
sh
w
a
te

r
su

m
m

e
r

b
lo

o
m

s
P
h
o
sp
h
a
te
+

N
od

u
la

ri
a

sp
u

m
ig

en
a

(n
o

d
u

la
ri

n
)

B
ra

ck
is

h
w
a
te

r
P

h
o

sp
h

a
te

T
ri

ch
od

es
m

iu
m

(‘
‘r

e
d

-t
id

e
’’

)
T

r

o
p

ic
a
l

F
e

fo
r

N
fi

x
a
ti

o
n
+

K
a
re

n
ia

br
ev

is
(N

S
P

,
fi

sh
k
il
ls

)
D

e
m

is
e

o
f
b
lo
o
m

R
iv

e
r

fl
o

w
F

e
,
M
n

V
it

a
m

in
s,

a
m

in
o

a
ci

d
s
+
+
K
a
re
n
ia

m
ik

im
ot

oi
(fi

sh
k
il
ls

)
P

o
ly

a
m

in
e
s

+
P

yc
n

o
cl

in
e
s

fr
o

n
ts

N
oc

ti
lu

ca
(‘

‘r
e
d

-t
id
e
’’
)
B
lo
o
m
w
in
d
o
w

In
d

ir
e
ct

;
h

ig
h

p
re

y
b

io
m

a
ss

P
h

a
g

o
tr

o
p

h
y

P
se

u
d
o-

n
it

zs
ch

ia
(A

S
P

)
?R

e
sp
o
n
d
s

to
ra

in
fa

ll
N

it
ra

te

P
fi

es
te

ri
a

a
C

ry

p
to

p
er

id
in

io
p
si

sa
C

h
a
n

g
e

in
n

u
tr

ie
n

t
re

g
im

e

C
ry

p
to

m
o

n
a
d

p
re

y
+

K
a
rl

od
in
iu
m

(

fi
sh

k
il

ls
)

C
ry
p
to
m
o
n
a
d
p
re

y
P

h
a
eo

cy
st

is
(g

e
la

ti
n

o
u

s
b

lo
o

m
s)

C
h
a
n
g
e
in
n
u
tr
ie
n
t
re
g
im
e

F
e
?

+

C
h
ry

so
ch

ro
m

u
li

n
a

(r
a
re

ly
fi

sh
k
il
ls

)
S

e
,

F
e
,

C
u

,
V

it
B

1
2

+
+

P
ry

m
n

es
iu

m
(fi

sh
k
il
ls

)
+

D
in

op

h
ys

is
(D

S
P
)
+
+
+

G
a
m

bi
er

d
is

cu
s

(

C
F

P
)
T
ro
p
ic
a
l

M
a
cr

o
a
lg

a
l

su
b

st
ra

te

T
a
x

a
id

e
n

ti
fi

e
d

a
s

re
sp

o
n

si
ve

to
te

m
p
e
ra
tu
re

,
la

n
d

ru
n

o
ff

,
n

u
tr
ie
n

ts
a
n

d
m

ix
e
d

-l
a
ye

r
d

e
p

th
,

a
n

d
p

h
ys
ic
a
l
tu
rb
u
le
n
ce

a
p

p
e
a
r

m
o

st
vu

ln
e
ra

b
le

to
cl

im
a
te

ch
a
n

g
e
.

‘‘
?’

’
in

d
i-

ca
te

s
li

m
it

e
d

e
vi

d
e
n

ce
.

P
S

P
,

p
a
ra

ly
ti

c
sh

e
ll

fi
sh
p
o

is
o

n
in

g
;

N
S

P
,

n
e
u

ro
to

x
ic

sh
e
ll

fi
sh
p
o
is
o
n
in
g
;

A
S

P
,
a
m

n
e
si

c
sh
e
ll
fi
sh
p
o
is
o
n
in
g
;

D
S

P
,

d
ia

rr
h

e
ti

c
sh
e
ll
fi
sh
p
o
is
o
n
in
g
;
C
F
P
,

ci
g

u
a
te

ra
fo

o
d

p
o
is
o
n
in

g
.

a
D

is
p

u
te

d
H

A
B

sp
e
ci

e
s.

224 G U S T A A F M . H A L L E G R A E F F

Most HAB species have been demonstrated to
have either some capability of mixotrophy ⁄ organic
nutrient uptake or a requirement for micronutrients
(Graneli and Turner 2006). Temperature plays a
crucial role in the bloom dynamics of the cyst-form-
ing PSP dinoflagellates and raphidophytes, as well
as for species such as G. catenatum, Noctiluca, and
many cyanobacteria, which have well-defined sea-
sonal temperature windows. However, once cells of
these species enter the water column, other factors
such as nutrients, turbulence, and grazing deter-
mine the outcome of competition. HAB species
show a perplexing diversity of biomass and toxicity
patterns (Cembella 2003), ranging from species
such as Dinophysis and Chrysochromulina, which can
cause toxicity problems even at very low cell concen-
trations, to species such as Phaeocystis and Noctiluca,
which are basically nontoxic but whose nuisance
value derives from their high biomass production.
Persistent near-monospecific algal blooms of, for
example, Aureococcus, Chrysochromulina, Prymnesium,
and Nodularia have recently been referred to as eco-
system disruptive algal blooms (EDABs), in which
toxic or unpalatable algal species disrupt grazing
and thus diminish nutrient supply via recycling
(Sunda et al. 2006). Could climate perturbations
perhaps create a niche for such HAB species?

From progress in the past three decades, it has
become abundantly clear that the niche of HAB spe-
cies is much wider than originally envisaged. HAB
species are not restricted to dinoflagellates but also
include diatoms, haptophytes, raphidophytes, and
cyanobacteria. Furthermore, they cover the com-
plete range from r-strategists (e.g., Pseudo-nitzschia,
Chattonella), whose success is due to their high
growth rates (r) and efficient use of nutrients, to
K-strategists (e.g., G. catenatum), which can achieve
high biomass levels by being energy (light) efficient,
for example, by vertical migration (Margalef 1978,
Smayda and Reynolds 2001). Taxa identified in
Table 1 as responsive to temperature, land runoff,
nutrients and mixed-layer depth, and physical turbu-
lence appear most vulnerable to climate change.
When and where climate-driven perturbations open
a new ‘‘niche,’’ any number of ecologically similar
organisms have the opportunity to emerge from the
background to become a HAB phenomenon.

Algal bloom range expansions and climate change. For
many HAB species, significant bloom episodes can
serve as stepping stones toward range expansions via
natural current systems, sometimes facilitated by
local climate events or ship ballast water dispersal.
The dinoflagellate Pyrodinium bahamense is presently
confined to tropical, mangrove-fringed coastal
waters of the Atlantic and Indo-West Pacific. A sur-
vey of cyst fossils (named Polysphaeridium zoharyii)
going back to the warmer Eocene 50 mya indicates
a much wider range of distribution in the past. For
example, in the Australasian region at present, the
alga is not found farther south than Papua New

Guinea, but some 120,000 years ago, the alga ran-
ged as far south as Bulahdelah (32�S) just north of
Sydney (McMinn 1988, 1989). There is concern
that, with increased warming of the oceans, this spe-
cies may return to Australian waters (Fig. 2). In the
tropical Atlantic, in areas such as Bahia Fosforescen-
te in Puerto Rico and Oyster Bay in Jamaica, the
bioluminescent blooms of Pyrodinium are a major
tourist attraction, but Pyrodinium blooms gained a
more sinister reputation in 1972 in Papua New Gui-
nea after red-brown water discolorations coincided
with the fatal food poisoning of three children in a
seaside village, diagnosed as PSP. Since then, the
incidence of toxic blooms has spread to Brunei and
Sabah (1976), the central (1983) and northern Phil-
ippines (1987), and Indonesia (North Mollucas).
Pyrodinium is a serious public health and economic
problem for tropical countries, all of which depend
heavily on seafood for protein. In the Philippines
alone, Pyrodinium has now been responsible for
>2,000 human illnesses and 100 deaths resulting
from the consumption of contaminated shellfish as
well as sardines and anchovies (Hallegraeff and
Maclean 1989). There exists circumstantial but
debated evidence of a coincidence between Pyrodi-
nium blooms and the ENSO (Maclean 1989, Azanza
and Taylor 2001). In the Pacific Basin, trade winds
and strong equatorial currents normally flow
westward, and cold upwelling occurs off Peru. In
contrast, during an ENSO event, trade winds are
weak, and anomalously warm equatorial water flows
eastward, and stratification is enhanced. Erickson

Fig. 2. Global distribution of Pyrodinium bahamense in recent
plankton (A) and much wider distribution in the fossil cyst
record (B) (after Hallegraeff 1993).

C L I M A T E C H A N G E A N D A L G A L B L O O M S 225

and Nishitani (1985) similarly reported exceptional
PSP episodes by Alexandrium tamarense ⁄ catenella in
the Pacific Northwest during seven out of nine
ENSO events between 1941 and 1984 (but see
Moore et al. 2008a for an alternative interpreta-
tion). An exceptional Karenia digitata red-tide event
in Hong Kong in 1998 (HK$250 M loss to aquacul-
ture) was associated with El Niño and altered ocean-
ographic conditions (Yin et al. 1999). In the North
Atlantic, the NAO reflects a north–south oscillation
in atmosphere mass between the Iceland-low and
the Azores high-pressure center. A high NAO means
increased westerly winds and milder temperatures
over northern Europe, and a low NOA causes cooler
temperatures due to decreased westerly winds. Bel-
grano et al. (1999) found significant correlations
between NAO, phytoplankton biomass, primary pro-
duction, and Dinophysis concentrations off Sweden.

Until recently, NSP by the dinoflagellate K. brevis
was considered to be endemic to the Gulf of Mexico
and the east coast of Florida, where red tides had
been reported as early as 1844. An unusual feature
of NSP is the formation of toxic aerosols by wave
action, which can lead to respiratory asthma-like
symptoms in humans. In 1987, a major Florida
bloom was dispersed by the Gulf Stream northward
into North Carolina waters, even though it has not
persisted there (Tester et al. 1991, 1993). Unexpect-
edly, in early 1993, >180 human NSPs were reported
from New Zealand. Most likely, this mixed bloom of
K. mikimotoi and related species was again triggered
by the unusual weather conditions at the time,
including higher than usual rainfall and lower than
usual temperature, which coincided with El Niño
(Rhodes et al. 1993, Chang et al. 1998).

Ciguatera caused by the benthic dinoflagellate
Gambierdiscus toxicus is a food-poisoning syndrome
caused by ingesting tropical fish and is well known
in coral reef areas in the Caribbean, Australia, and
especially French Polynesia (Fig. 3). Whereas, in a
strict sense, this is a completely natural phenome-
non, from being a rare disease two centuries ago,

ciguatera has now reached epidemic proportions in
French Polynesia. From 1960 to 1984, >24,000
patients were reported from this area, which is more
than six times the average for the Pacific as a whole
(Bagnis et al. 1985). Evidence is accumulating that
reef disturbance by hurricanes, military and tourist
developments, as well as coral bleaching (linked to
global warming), increased water temperatures
(>29�C preferred in culture), and perhaps in future
increasing coral damage due to ocean acidification
(Hoegh-Guldberg 1999) are increasing the risk of
ciguatera by freeing up space for macroalgae for
Gambierdiscus to colonize upon. During El Niño
events, ciguatera increased on Pacific islands where
sea surface temperatures increased (Hales et al.
2001). In the Australian region, G. toxicus is well
known from the tropical Great Barrier Reef and
southward to just north of Brisbane (25�S), but in
the past 5 years, this species has undergone an
apparent range expansion into southeast Australian
seagrass beds as far south as Merimbula (37�S),
aided by a strengthening of the East Australian
Current (S. Brett, M. de Salas, and G. Hallegraeff,
unpublished data). A similar expansion of Gambier-
discus into the Mediterranean and eastern Atlantic
has been reported (Aligizaki et al. 2008), and
blooms of the associated benthic dinoflagellate
genus Ostreopsis are also an increasingly common
phenomenon in temperate regions worldwide
(Shears and Ross 2009).

In the same Australian region, the red-tide dino-
flagellate Noctiluca scintillans (known from Sydney
as early as 1860) has expanded its range from Syd-
ney into Southern Tasmanian waters since 1994
where it has caused problems for the salmonid fish
farm industry (Fig. 4). In the North Sea, an analo-
gous northward shift of warm-water phytoplankton

Fig. 3. Current global distribution of ciguatera food poison-
ing from fish (after Hallegraeff 1993).

Fig. 4. Apparent range expansion of Noctiluca scintillans in the
Australian region, comparing distribution records in 1860–1950,
1980–1993 (expansion of blooms in the Sydney region),
1994–2005 (range extension into Tasmania), and 2008 (first
reports in Queensland, West Australia, and South Australia).
After Hallegraeff et al. (2008).

226 G U S T A A F M . H A L L E G R A E F F

has occurred due to regional climate warming
(Edwards and Richardson 2004, Richardson and
Schoeman 2004). For example, Ceratium trichoceros,
previously found only south of the British Isles, has
expanded its geographic range to the west coast of
Scotland and the North Sea, and the subtropical
Ceratium hexacanthum moved 1,000 km northward in
40 years (Hays et al. 2005). At the same time, Proro-
centrum, Ceratium furca, and Dinophysis increased
along the Norwegian coast, and Noctiluca increased
in the southern North Sea (Fig. 5). It is difficult to
untangle the role of climate change and eutrophi-
cation in some of these species patterns. Dale
(2009) used the dinoflagellate cyst record from the
last 100 years to discriminate between the role of
local eutrophication events within the Skagerrak
(indicated by a shift to heterotrophic dinocysts,
reflecting increased diatom prey) and the role of
regional variation in the NAO, thought to have
increased transport of relatively nutrient-rich North
Sea water into the system (indicated by increased
Lingulodinium benefitting from added nutrients dur-
ing warm summers and increased stratification). In
the same Skagerrak area, Thorsen et al. (1995),
working with a much longer sediment core, docu-
mented the immigration into the region �6000
B.P. (before present) of Gymnodinium nolleri (ini-
tially reported as G. catenatum), which achieved
bloom proportions from 2000 to 500 B.P. during a
warming period, followed by a near extinction dur-
ing the cooling period that commenced in 300 B.P.
(Fig. 6). Reduction between 1997 and 2002 of Arc-
tic ice by 33% allowed greater flow through the
trans-Arctic pathway from the Pacific to the Atlantic
and was associated with the first appearance in
800,000 years in the Labrador Sea of the diatom
Neodenticula seminae (Reid et al. 2007). While warm-
water species can be expected to expand their dis-
tribution, cold-water species will contract their
range (compare Beaugrand et al. 2002 for zoo-
plankton). For example, the cold-water dinoflagel-
late cyst Bitectatodinium tepikiense currently is
confined to Tasmania (43�S), but in the last inter-
glacial period (120,000 years ago), it was found as
far north as Sydney (34�S) (McMinn and Sun

1994). On top of range extensions driven by grad-
ual climate change, ship ballast water translocations
continue to alter species distributions. The two
mechanisms interact since ecosystems disturbed by
pollution or climate change are more prone to
ballast water invasions (Stachowicz et al. 2002).
Similarly, melting of Arctic sea ice and opening of
new Arctic shipping lanes will encourage range
expansions via both natural current dispersal as
well as ballast water invasions.

Fig. 5. Decadal anomaly maps (difference between long-term 1960–1989 mean and the 1990–2002 period) for four common HAB spe-
cies (from left to right): Prorocentrum, Ceratium furca, Dinophysis, and Noctiluca in the North Atlantic. Note the increase in Prorocentrum, C.
furca, and Dinophysis along the Norwegian coast, and increase in Noctiluca in the southern North Sea, reportedly associated with a contraction
of the Subpolar Gyre to the west allowing subtropical water to penetrate farther north (adapted from Edwards et al. 2008, with permission).

Fig. 6. Quantitative distribution of Gymnodinium nolleri cysts
and total dinoflagellate cysts (cysts Æ g)1 dry sediment) in an
860 cm long sediment core from the southern Kattegat (after
Thorsen et al. 1995).

C L I M A T E C H A N G E A N D A L G A L B L O O M S 227

Impact of global warming and sea surface temperature
change. Phytoplankton grow over a range of tem-
peratures characteristic of their habitat, and growth
rates are usually higher at higher temperature, but
considerably lower beyond an optimal temperature
(Eppley 1972). Natural populations of phytoplank-
ton often occur at temperatures suboptimal for pho-
tosynthesis, and it is believed that this distribution is
designed to avoid risking abrupt declines in growth
associated with the abrupt incidence of warmer tem-
peratures (Li 1980). Temperature effects on phyto-
plankton growth and composition are more
important in shallow coastal waters, which experi-
ence larger temperature fluctuations than oceanic
waters. Predicted increasing sea surface tempera-
tures of 2�C–4�C may shift the community composi-
tion toward species adapted to warmer temperatures
as observed in the temperate North Atlantic
(Edwards and Richardson 2004). Several well-stud-
ied PSP dinoflagellates, such as A. catenella in Puget
Sound (Moore et al. 2008b) and G. catenatum in
Tasmania, Australia (Hallegraeff et al. 1995), bloom
in well-defined seasonal temperature windows
(>13�C and >10�C, respectively). Climate change
scenarios are predicted to generate longer-lasting
bloom windows (Fig. 7).

In the North Sea, the NAO has been shown to
affect the length of the phytoplankton growing sea-
son, which has increased in parallel with the warm-
ing of sea surface temperatures (Barton et al. 2003).
Seasonal timing of phytoplankton blooms is now
occurring there up to 4–6 weeks earlier (Fig. 8).
However, where individual zooplankton or fish graz-
ers are differentially impacted by ocean warming,
this may have cascading impacts on the structure of
marine food webs (‘‘match-mismatch’’ sensu Cush-

ing 1974). Replacement in the North Sea of the
cold-water copepod Calanus finmarchicus by the
warm-water Calanus helgolandicus has been circum-
stantially associated with the decline of cod
(Edwards et al. 2008). Similarly, Attrill et al. (2007)
predict an increasing occurrence of jellyfish in the
central North Sea over the next 100 years related to
the increased Atlantic inflow to the northern North
Sea.

Sea-level rise, wind, and mixed-layer depth. Increasing
sea surface temperature and water column stratifica-
tion (shallowing of the mixed layer) can be
expected to have a strong impact on phytoplankton
because of the resource requirements and tempera-
ture ranges that species are adapted to. Wind deter-
mines the incidence of upwelling and downwelling,
which in turn strongly affect the supply of macronu-
trients to the surface (recognized as drivers of
G. catenatum blooms off Spain; Fraga and Bakun
1990). Climate change may thus affect the timing
and strength of coastal upwellings. Broad changes
in ocean circulation such as those comprising the
deep-ocean conveyor belt (Rahmstorf 2002) can also
cause displacements to current systems and associ-
ated algal bloom phenomena. Wind-driven currents
can transport phytoplankton away from a region
and affect the size and frequency of formation of
mesoscale features such as fronts and eddies.
Locally, wind intensity strongly influences depth
and intensity of vertical mixing in the surface layer,
thereby affecting phytoplankton access to nutrients,
light availability for algal photosynthesis, and phyto-
plankton exposure to potentially harmful UVB radi-
ation. Winds can also influence the supply of iron
to the surface ocean through aeolian transport of
dust from land to sea, contributing micronutrients
such as iron, which has been shown to stimulate
K. brevis blooms off Florida (Walsh and Steidinger
2001). Extreme climate events such as hurricanes

Fig. 8. Long-term monthly values of ‘‘phytoplankton color’’
in the central North Sea from 1948 to 2001. Circles denote >2 SD
above the long-term monthly mean and a major regime shift
�1998. Note an apparent shift toward earlier spring and autumn
phytoplankton blooms (after Edwards 2004, with permission).

Fig. 7. Scenarios for warmer sea surface temperature condi-
tions in Puget Sound by 2, 4, and 6�C would widen the >13�C
window (in gray) of accelerated growth for the PSP dinoflagellate
Alexandrium catenella. After Moore et al. (2008b). PSP, paralytic
shellfish poisoning.

228 G U S T A A F M . H A L L E G R A E F F

are also known to expand the existing distribution
of cyst-producing toxic dinoflagellates (e.g., A. tama-
rense in New England after a 1972 hurricane; Ander-
son 1997). Sea-level rises of 18 to 59 cm (up to
1 m) predicted by 2100 (IPCC 2008) have the
potential to increase the extent of continental shelf
areas, providing shallow, stable water columns favor-
ing phytoplankton growth. The proliferation of coc-
colithophorids in the Cretaceous geological period
has been attributed to expanding continental shelf
areas (Bown et al. 2004).

Finally, increasing temperature driven by climate
change is predicted to lead to enhanced surface
stratification, more rapid depletion of surface nutri-
ents, and a decrease in replenishment from deep
nutrient-rich waters (Fig. 9A). This in turn will lead
to changes in phytoplankton species, with smaller
nano- and picoplankton cells with higher surface
area:volume ratios (better able to cope with low
nutrient levels) favored over larger cells. A decline
in silica concentrations is widely expected in a
warming world (reported, e.g., from the Mediterra-
nean; Goffart et al. 2002), and this is anticipated to
restrict the abundance of diatoms. Mixing depth
affects sea surface temperature, the supply of light
(from above) and nutrients (from below), and phy-
toplankton sinking losses within the surface layer.
Climate models predict changes in mixed-layer
depth in response to global warming for large
regions of the global ocean. In the North Pacific,
decadal-scale climate and mixed-layer variability,
(Hayward 1997) and, in the North Atlantic, longer-
term changes in wind intensity and stratification
since the 1950s have been associated with consider-
able changes in phytoplankton community structure
(Richardson and Schoeman 2004). In regions with
intermediate mixing depth, increased stratification

is expected to result in decreased phytoplankton
biomass due to reductions in nutrient supply. The
observed reductions in open-ocean productivity
(‘‘desertification’’) during ENSO warming events
provide insight on how future climate change can
alter marine food webs (Behrenfeld et al. 2006).
Conversely, in high-latitude regions with relatively
deep mixing and nonlimiting nutrients, decreasing
mixing depth is expected to result in higher phyto-
plankton biomass because of increased light avail-
ability (Fig. 9B; Doney 2006).

Impact of heavy precipitation and storm events and flash
floods. Episodic storm events affect the timing of
freshwater flow, residence time, and magnitude and
time of nutrient pulses. Changes in the amount or
timing of rainfall and river runoff affect the salinity
of estuaries and coastal waters. Salinity is relatively
constant throughout the year in most oceanic waters
and in coastal areas that receive little freshwater
input. Coastal phytoplankton is subject to more vari-
ation in salinity than phytoplankton in oceanic
waters. While some species grow well over a wide
range of salinities, other species grow best only at
salinities that are low (estuarine), intermediate
(coastal), or high (oceanic species). Freshwater also
modifies the stratification of the water column,
thereby affecting nutrient resupply from below.
While diatoms seem to be negatively affected by the
inhibition of mixing associated with river discharge,
dinoflagellates often benefit as this usually increases
stratification and the availability of humic substances
for growth (Graneli and Moreira 1990, Doblin et al.
2005). PSP dinoflagellate blooms of G. catenatum (in
Tasmania; Hallegraeff et al. 1995) and A. tamarense
(off Massachusetts; Anderson 1997) tend to be clo-
sely associated with land runoff events. In Hiro-
shima Bay, blooms of the fish-killing raphidophyte

Fig. 9. Predicted phytoplank-
ton response to increased tem-
perature in ocean surface waters:
(A) reduced productivity in the
thermally stratified water of tropi-
cal and midlatitudes caused by
reduced nutrient supply; (B)
increased productivity at polar
and subpolar salt-stratified oceans
where reduced mixing keeps
plankton closer to the well-lit
nutrient-sufficient surface layers.
Adapted from Doney (2006).

C L I M A T E C H A N G E A N D A L G A L B L O O M S 229

Chattonella marina followed typhoon-induced accre-
tion of nutrient-rich land runoff (Kimura et al.
1973). Climate change is predicted to cause rainfall
to occur in more concentrated bursts followed by
long dry periods, thus favoring dinoflagellates.

Increased CO2 and ocean acidification. Increasing
atmospheric CO2 is leading to ocean acidification,
which could potentially have an adverse impact on
calcifying organisms, the most important of which
in terms of biomass and carbon sequestration is the
coccolithophorid Emiliania huxleyi (Riebesell et al.
2000). Calculations based on CO2 measurements of
the surface oceans indicate that uptake by the
oceans of approximately half the CO2 produced by
burning fossil fuels has already led to a reduction of
surface pH by 0.1 unit. Under the current scenario
of continuing global CO2 emissions from human
activities, average ocean pH is predicted to fall by
0.4 units by the year 2100 (Orr et al. 2005). Such
pH is lower than has been experienced for millen-
nia, and, critically, this rate of change is 100 times
faster than ever experienced in the known history
of our planet (Raven et al. 2005b). Experimental
manipulations of pH in E. huxleyi cultures have both
produced reduced (Riebesell et al. 2000) and
enhanced calcification and growth (Iglesias-
Rodriguez et al. 2008). Strikingly, depression of
calcification under high CO2 is most expressed
under high-light conditions (Feng et al. 2008;
Fig. 10), emphasizing the need to carefully consider
factor interactions in ecophysiological experiments.
Decreasing pH < 8.0 has been observed to nega- tively affect nitrification in marine bacteria and therefore could potentially reduce nitrate availability for plankton algae. The nitrogen-fixing tropical cya- nobacterium Trichodesmium may be a beneficiary of ocean acidification, however (Hutchins et al. 2007). Decreasing pH has also been found to increase the availability of toxic trace elements such as copper. As the relative consumption of HCO3

) and CO2
differs between phytoplankton species, changes in
their availability may affect phytoplankton at the cel-
lular, population, and community levels. Most HAB
species tested thus far lack carbon concentrating
mechanisms (CCMs), and hence their photosyn-
thetic performance may benefit from increased
atmospheric CO2 (e.g., E. huxleyi in Fig. 11). How-
ever, photosynthetic performance in diatom species
such as Skeletonema, for which photosynthesis is
already CO2 saturated, will remain constant (Beardall
and Raven 2004). In bioassays in the equatorial
Pacific, high CO2 (750 ppm) favored the hapto-
phyte Phaeocystis at the expense of diatoms, whereas
at low CO2 (150 ppm), diatom growth was stimu-
lated (Tortell et al. 2002). Riebesell et al. (2007),
working with mesocosms dominated by diatoms and
coccolithophorids, observed increases in productiv-
ity of 27% and 39% when CO2 levels were elevated
to 700 and 1,050 ppm, respectively. Similarly, Schip-
pers et al. (2004) predict in nutrient-replete systems

a 10%–40% increase of marine productivity for spe-
cies with low bicarbonate affinity, thus potentially
aggravating some coastal algal blooms.

UV radiation. Although the implementation of
the Montreal Protocol has done much to slow the
build-up of chlorofluorocarbons in the stratosphere,
elevated UVB levels from the Arctic and Antarctic

Fig. 10. Calcification of the coccolithophorid Emiliania huxleyi
(expressed as the particulate inorganic carbon [PIC] to particu-
late organic carbon ratio [POC]) as a function pCO2 (ambient
375 ppm or high 750 ppm), temperature (ambient 20�C or high
24�C), and light (low = black bars; high = open bars). Neither
pCO2, light, nor temperature influences PIC:POC under low
light, but calcification is most reduced under a combination of
high light · high pCO2 (after Feng et al. 2008).

Fig. 11. Photosynthesis of three different phytoplankton spe-
cies (diatom Skeletonema costatum, haptophytes Phaeocystis globosa
and Emiliania huxleyi) with respect to CO2 sensitivity (adapted
from Rost and Riebesell 2004). Microalgal species differ in their
responses to CO2, which implies that a high-CO2 ocean will
induce shifts in phytoplankton species composition.

230 G U S T A A F M . H A L L E G R A E F F

ozone holes are expected to persist until at least
2050. UVB can negatively affect several physiological
processes and cellular structures of phytoplankton,
including photosynthesis, nutrient uptake, cell
motility and orientation, algal life span, and DNA
(Häder et al. 1991). Whereas shorter wavelengths
generally cause greater damage per dose, inhibition
of photosynthesis by ambient UVB increases linearly
with increasing total dose. In clear oceanic waters,
UVB radiation can reach depths of at least 30 m.
Although some phytoplankton may acclimate to,
compensate for, or repair damage by UVB, this
involves metabolic costs, thereby reducing energy
available for cell growth and division. Raven et al.
(2005a) suggest that UVB intensity affects the size
ratio in phytoplankton communities because small
cells are more prone to UVB and have compara-
tively high metabolic costs to screen out damaging
UVB. Many surface-dwelling red-tide species of raph-
idophytes and dinoflagellates possess UVB-screening
pigments, which give them a competitive advantage
over species lacking such UV protection (Jeffrey
et al. 1999). In some species, nutrient limitation of
either N or P (from increased water column stratifi-
cation) can enhance the sensitivity of cells to UVB
damage (Shelly et al. 2002).

Feedback mechanisms. One cannot talk about the
impact of climate on phytoplankton without also
considering the impact of phytoplankton on
climate (Fig. 12). Phytoplankton play a key role in
several global biogeochemical cycles and thereby
exert important feedback effects on climate by
influencing the partitioning of climate-relevant
gases between the ocean and the atmosphere.
Some species (e.g., Emiliania, Phaeocystis) are pro-
ducers of dimethylsulfonium propionate, a precur-
sor of dimethylsulfoxide (DMS), which in the
atmosphere is oxidized into sulfate, which forms
condensation nuclei for clouds (Charleson et al.
1987). Subsequent work on DMS has clarified that
it is not just phytoplankton, however, but also zoo-
plankton and bacterial food-web structure and
dynamics that drive oceanic production of atmo-
spheric sulfur. Phytoplankton, therefore, indirectly
affect albedo and precipitation and hence coastal
runoff, salinity, water column stratification, and
nutrient supply.

Through the process of photosynthesis, phyto-
plankton constitute a major consumer of CO2.
The ability of the oceans to act as a sink for
anthropogenic CO2 largely relies on the conver-
sion of this gas by phytoplankton into particulate

Fig. 12. Summary diagram of known feedback mechanisms between physicochemical climate variables and biological properties of
marine phytoplankton systems. Red: Greenhouse warming raises surface temperatures and causes a shoaling of mixed-layer depths but can
also have broader impacts on global currents, upwelling, and even the deep-ocean conveyor belt. Blue: Increased atmospheric CO2 drives
the biological pump, can alter phytoplankton species composition, and can alter ocean pH, influencing calcification of coccolithophorids
but also nutrient availability. Green: Nutrient impacts from water column stratification, as well as linked to shifts in marine food-web struc-
ture, influenced by fishing, eutrophication, and even ship ballast water invasions. Yellow: Marine food-web structure, including top-down
as well as bottom-up influences on phytoplankton species composition. Orange: Selected phytoplankton such as coccolithophorids pro-
duce dimethylsulfoxide (DMS), acting as cloud condensation nuclei, thereby reducing solar irradiation. Other anthropogenic influences
in terms of eutrophication, shipping (ballast water introductions), and fishing are also indicated. Without exception, all perturbations will
drive changes in phytoplankton species composition.

C L I M A T E C H A N G E A N D A L G A L B L O O M S 231

organic matter and subsequent partial loss to the
deep ocean (so-called biological pump). Any
reduction in net ocean CO2 uptake caused by
shifts in ocean circulation or reduced phytoplank-
ton growth in surface waters reducing the export
of organic matter to the deep sea via the biologi-
cal pump could lead to an acceleration in the
rate of atmospheric CO2 increase and global
warming. Models have estimated that a 50%
decrease in oceanic calcification from ocean acidi-
fication thus would reduce atmospheric CO2 by
10–40 ppm (Heinze 2004, Munhoven 2007), equiv-
alent to 5–20 years of industrial emissions. Con-
versely, an increase in calcification would increase
CO2 levels by a similar amount (carbonate count-
erpump). Coupled climate-carbon models are
increasingly revealing feedback mechanisms, which
were completely unpredicted from first principles.
Stratosphere ozone depletion increases the
strength of Southern Ocean winds and thereby
the ventilation of carbon-rich deep water, with
consequences of reduced ocean carbon uptake
and enhanced ocean acidification (Lenton et al.
2009).

Woods and Barkmann’s (1993) ‘‘plankton
multiplier’’ is an example of a positive feedback
mechanism linking greenhouse warming to the
biological pump. Enhanced greenhouse CO2
induces ocean surface warming, diminishing winter
convection and nutrient availability and thereby
primary production, thus weakening the biological
pump and further enhancing atmospheric CO2.
These authors suggested that a similar mechanism
may have underpinned global warming at the end
of the ice ages when the Milankovich effect
enhanced greenhouse warming. Surface phyto-
plankton blooms influence the oceanic heat bud-
get, and this is dependent not only on the chl
biomass but also on its precise vertical distribution
(Frouin and Lacobellis 2002).

Finally, another powerful mechanism for algal
bloom formation occurs through ‘‘top-down con-
trol’’ of the marine food web (Turner and Grane-
li 2006). Overfishing removes top fish predators,
stimulating small fish stocks, which graze away
zooplankton, thus relieving phytoplankton grazing
pressure. Differential impacts of climate change
on individual zooplankton or fish grazers (uncou-
pling between trophic levels) thus can result in
stimulation of HABs. Figure 12 summarizes known
feedback mechanisms between physicochemical cli-
mate variables and biological properties of marine
phytoplankton systems, altogether confronting us
with a formidable predictive challenge. Without
exception, all physicochemical climate stressors
drive changes in phytoplankton species composi-
tion, but the precise direction of such changes
(i.e., whether they may lead to HABs) remains lar-
gely unpredictable in view of our current incom-
plete knowledge of phytoplankton ecophysiology.

CONCLUSIONS

Climate change confronts marine ecosystems with
multifactorial stressors, such as increased tempera-
ture, enhanced surface stratification, alteration of
ocean currents, intensification or weakening of
nutrient upwelling, stimulation of photosynthesis by
elevated CO2, reduced calcification from ocean acid-
ification, and changes in land runoff and micronu-
trient availability. Complex factor interactions are
rarely covered by simulated ecophysiological experi-
ments, and the full genetic diversity and physiologi-
cal plasticity of phytoplankton taxa are rarely
considered. Traditional experimental challenges last
days to weeks and impose new growth conditions
rather quickly, thus only allowing for limited accli-
mation (testing short-term physiological plasticity
but without genetic changes). Predicted global
change will occur gradually over decades, allowing
for adaptation of species to perhaps become geneti-
cally and phenotypically different from the present
population. Laboratory studies should aim to mimic
environmental conditions as closely as possible
(Rost et al. 2008). A typical example is the problem
of the potential impact of increased CO2 on the
coccolithophorid E. huxleyi. Initial concerns focused
on reduced calcification (Riebesell et al. 2000), but
we now recognize that increased CO2 at the same
time stimulates photosynthesis (Iglesias-Rodriguez
et al. 2008). Complex factor interactions between
increased CO2, light, and temperature on the calci-
fication versus photosynthesis dynamics of E. huxleyi
have been demonstrated by Feng et al. (2008), while
geographic strain variability of this ‘‘cosmopolitan’’
taxon has confounded the extensive literature on
this taxon (Langer et al. 2009). At the same time,
field observations of E. huxleyi are suggesting an
apparent range expansion in the past two decades
toward both the Arctic (Bering Sea; Merico et al.
2003) and Antarctic (Cubillos et al. 2007), but the
environmental drivers underpinning this are by no
means clear. Undoubtedly, there will be winners
and losers from climate change, and one thing we
can be certain about is local changes in species
composition, abundance, and timing of algal
blooms.

The greatest problems for human society will be
caused by being unprepared for significant range
extensions of HAB species or the increase of algal
biotoxin problems in currently poorly monitored
areas. While, for example, ciguatera contamination
would be expected and monitored for in tropical
coral reef fish, with the apparent range extension of
the causative benthic dinoflagellate into warm-tem-
perate seagrass beds of Southern Australia, other
coastal fisheries unexpectedly could be at risk.
Range expansion of Noctiluca from Sydney to Tasma-
nia exposed the salmonid aquaculture industry to a
novel HAB problem. Polar expansion of domoic-
acid-producing Pseudo-nitzschia australis could pose a

232 G U S T A A F M . H A L L E G R A E F F

novel threat to krill-feeding whales (Lefebvre et al.
2002). Similarly, incidences of increased surface
stratification in estuaries or heavy precipitation or
extreme storm events are all warning signs that call
for increased vigilance of monitoring seafood prod-
ucts for algal biotoxins even in areas not currently
considered to be at risk. Changes in phytoplankton
communities provide a sensitive early warning for
climate-driven perturbations to marine ecosystems.

Only with improved global ocean observation sys-
tems (GOOS) can we hope to quantitatively monitor
the key variables identified in this review. New,
improved, and expanded ocean sensor capabilities
(e.g., argo floats, ocean gliders, coastal moorings
and coastal radar, multiwavelength and variable flu-
orometers, optical sensors) are necessary to realize
the full potential of in situ ocean-observing net-
works in support of integrated satellite-derived
‘‘ocean color’’ maps and expanded biological and
biogeochemical observations (continuous plankton
recorder, ecogenomics). Observations from the indi-
vidual components of such systems must be inte-
grated through data management and
communication capabilities that provide open
searchable access and routine delivery to all users.
Sustained observations, process research, and mod-
eling should determine fluxes and cycling of bio-
geochemical variables, identify impacts on
ecosystems, and resolve feedback from ecosystems
on climate. Achieving this will require extensive
infrastructure investment and poses a major chal-
lenge for the marine science community. It is pleas-
ing to see that a number of national (e.g., the U.S.
NSTC Joint Subcommittee on Ocean Science and
Technology 2007 Ocean Observatories Initiative
[OOI], the Australian Integrated Marine Observing
System [IMOS 2009]) and international programs
(e.g., the Intergovernmental Oceanographic Com-
mission of UNESCO’s GEOHAB) are actively pursu-
ing these ambitious goals, but necessary if we wish
to define management options, forecast ocean-
related risks to human health and safety, and shed
light on the impact of climate variability on marine
life and humans in general.

An earlier version appeared as a section of FAO Assessment
and Management of Fish Safety and Quality Technical Paper, and I
am grateful to Dr. Iddya Karunasagar for inviting me to con-
tribute to that effort. I thank my Tasmanian colleagues Prof.
Andrew McMinn, Prof. Harvey Marchant, and Prof. Tom
Trull for valuable discussions on the ocean carbon pump;
Prof. Chris Reid and Dr. Anthony Richardson (Sir Alister
Hardy Foundation for Ocean Science, Plymouth) for insights
into the Continuous Plankton Recorder program; Prof. Barrie
Dale for discussions on the fossil dinoflagellate cyst record;
and Dr. Stephanie Moore (University of Washington) for clar-
ifying the intricacies of ENSO, PDO, and NAO.

Aligizaki, K., Nikolaidis, G. & Fraga, S. 2008. Is Gambierdiscus
expanding to new areas? Harmful Algae 36:6–7.

Andersen, R. A. [Ed.] 2005. Algal Culturing Techniques. Elsevier, New
York, 596 pp.

Anderson, D. M. 1997. Bloom dynamics of toxic Alexandrium species
in the northeastern U.S. Limnol. Oceanogr. 42:1009–22.

Anderson, D. M., Cembella, A. D. & Hallegraeff, G. M. [Eds.] 1998.
Physiological Ecology of Harmful Algal Blooms. Proc. NATO-ASI
Workshop, Bermuda. Springer Verlag, Heidelberg, Germany,
662 pp.

Attrill, M., Wright, J. & Edwards, M. 2007. Climate-related increases
in jellyfish frequency suggest a more gelatinous future for the
North Sea. Limnol. Oceanogr. 52:480–5.

Azanza, R. V. & Taylor, F. J. R. 2001. Are Pyrodinium blooms in
Southeast Asia recurring and spreading? A view at the end of
the millennium. Ambio 30:356–64.

Bagnis, R., Bennett, J. & Barsinas, M. 1985. Epidemiology of
ciguatera in French Polynesia from 1960 to 1984. In Gabrie, C.
& Salvat, B. [Eds.] Proc. 5th Int. Coral Reef Congress. Antenne
Museum, Ephe Moorea, Tahiti, pp. 475–82.

Barton, A. D., Greene, C. H., Monger, B. C. & Pershing, A. J. 2003.
Continuous plankton recorder survey of phytoplankton mea-
surements and the North Atlantic Oscillation: interannual to
multidecadal variability. Prog. Oceanogr. 58:337–58.

Bates, S. S., Garrison, D. L. & Horner, R. A. 1998. Bloom dynamics
and physiology of domoic-acid producing Pseudo-nitzschia spe-
cies. In Anderson, D. M., Cembella, A. D. & Hallegraeff, G. M.
[Eds.] Physiological Ecology of Harmful Algal Blooms. NATO ASI
Series, Vol. G41. Springer Verlag, Heidelberg, Germany, pp.
267–92.

Beardall, J. & Raven, J. A. 2004. The potential effects of global
climate change on microalgal photosynthesis, growth and
ecology. Phycologia 43:26–41.

Beardall, J. & Stojkovic, S. 2006. Microalgae under global environ-
mental change: implications for growth and productivity,
populations and trophic flow. Science Asia 32(Suppl. 1):1–10.

Beaugrand, G., Reid, P. C., Ibanez, F., Lindley, J. A. & Edwards, M.
2002. Reorganization of North Atlantic marine copepod bio-
diversity and climate. Science 296:1692–4.

Behrenfeld, M. J., O’Malley, R. T., Siegel, D. A., McClain, C. R.,
Sarmiento, J. L., Feldman, G. C., Milligan, A. J., Falkowski, P. G.,
Letelier, R. M. & Boss, E. S. 2006. Climate-driven trends in
contemporary ocean productivity. Nature 444:752–5.

Belgrano, A., Lindahl, O. & Hernroth, B. 1999. North Atlantic
Oscillation, primary productivity and toxic phytoplankton in
the Gullmar Fjord, Sweden (1985–1996). Proc. R. Soc. Lond. B
Biol. Sci. 266:425–30.

Bown, P. R., Lees, J. A. & Young, J. R. 2004. Calcareous nano-
plankton evolution and diversity through time. In Thierstein,
H. R. & Young, J. R. [Eds.] Coccolithophores: From Molecular
Processes to Global Impact. Springer Verlag, Heidelberg, Ger-
many, pp. 481–508.

Burkholder, J. M. & Glibert, P. M. 2006. Intraspecific variability: an
important consideration in forming generalizations about
toxigenic algal species. Afr. J. Mar Sci. 28:177–80.

Cembella, A. D. 2003. Chemical ecology of eukaryotic microalgae
in marine ecosystems. Phycologia 42:420–47.

Chang, F. H., Sharples, J., Grieve, J. M., Miles, M. & Till, D. G. 1998.
Distribution of Gymnodinium cf. breve and shellfish toxicity
from 1993 to 1995 in Hauraki Gulf, New Zealand. In Reguera,
B., Blanco, J., Fernandez, M. L. & Wyatt, T. [Eds.] Harmful
Algae. Xunta de Galicia and IOC of UNESCO, Grafisant, San-
tiago de Compostela, pp. 139–42.

Charleson, R. J., Lovelock, J. E., Andreae, M. O. & Warren, S. G.
1987. Oceanic phytoplankton, atmospheric sulphur, cloud al-
bedo and climate. Nature 326:655–61.

Collins, S. & Bell, G. 2004. Phenotypic consequences of 1000 gen-
erations of selection at elevated CO2 in a green alga. Nature
431:566–9.

Cubillos, J. C., Wright, S. W., Nash, G., de Salas, M. F., Griffiths, B.,
Tilbrook, B., Poisson, A. & Hallegraeff, G. M. 2007. Calcifica-
tion morphotypes of the coccolithophorid Emiliania huxleyi in
the Southern Ocean: change in 2001–2006 compared to his-
toric data. Mar. Ecol. Prog. Ser. 384:47–54.

Cushing, D. H. 1974. The natural regulation of fish populations. In
Harden-Jones, F. R. [Ed.] Sea Fisheries Research. Elek Science,
London, pp. 399–412.

C L I M A T E C H A N G E A N D A L G A L B L O O M S 233

Dale, B. 2001. The sedimentary record of dinoflagellate cysts:
looking back into the future of phytoplankton blooms. Sci.
Mar. 65:257–72.

Dale, B. 2009. Eutrophication signals in the sedimentary record of
dinoflagellate cysts in coastal waters. J. Sea Res. 61:103–13.

Dale, B. & Yentsch, C. M. 1978. Red tide and paralytic shellfish
poisoning. Oceanus 2:41–9.

Doblin, M., Thompson, P. A., Revill, A. T., Butler, E. C. V., Black-
burn, S. I. & Hallegraeff, G. M. 2005. Vertical migration of the
toxic dinoflagellate Gymnodinium catenatum under different
concentrations of nutrients and humic substances in culture.
Harmful Algae 5:665–77.

Doney, S. C. 2006. Plankton in a warmer world. Nature 444:695–6.
Edwards, M. 2004. Phytoplankton blooms in the North Atlantic:

results from the Continuous Plankton Recorder survey
2001 ⁄ 2002. Harmful Algae News 25:1–3.

Edwards, M., Johns, D. G., Beaugrand, G., Licandro, P., John, A. W.
G. & Stevens, D. P. 2008. Ecological status report: results from
the CPR survey 2006 ⁄ 2007. SAHFOS Tech. Rep. 5:1–8.

Edwards, M. & Richardson, A. J. 2004. The impact of climate
change on the phenology of the plankton community and
trophic mismatch. Nature 430:881–4.

Eppley, R. W. 1972. Temperature and phytoplankton growth in the
sea. Fish. Bull. 70:1063–85.

Erickson, G. & Nishitani, L. 1985. The possible relationship of El
Nino ⁄ Southern Oscillation events to interannual variation in
Gonyaulax populations as shown by records of shellfish toxicity.
In Wooster, W. S. & Flaherty, D. L. [Eds.] El Niño North: Niño
Effects in the Eastern Subarctic Pacific. Washington Seagrant
Program, University of Washington, Seattle, pp. 283–90.

Falkowski, P. G. & Oliver, M. J. 2007. Mix and match: how climate
selects phytoplankton. Nat. Rev. 5:813–9.

Feng, Y., Warner, M. E., Zhang, Y., Sun, J., Fu, F.-X., Rose, J. M. &
Hutchins, D. A. 2008. Interactive effects of increased pCO2,
temperature and irradiance on the marine coccolithophore
Emiliania huxleyi (Prymnesiophyceae). Eur. J. Phycol. 43:87–98.

Fraga, S. & Bakun, A. 1990. Global climate change and harmful
algal blooms: the example of Gymnodinium catenatum on the
Galician coast. In Smayda, T. J. & Shimizu, Y. [Eds.] Toxic
phytoplankton blooms in the sea. Dev. Mar. Biol. 3:59–65.

Frouin, R. & Lacobellis, S. F. 2002. Influence of phytoplankton on
the global radiation budget. J. Geographic Res. 107:5-1–-10.

Gentien, P. 1998. Bloom dynamics and ecophysiology of the
Gymnodinium mikimotoi species complex. In Anderson, D. M.,
Cembella, A. D. & Hallegraeff, G. M. [Eds.] Physiological Ecology
of Harmful Algal Blooms. NATO ASI Series, Vol. G41. Springer
Verlag, Heidelberg, Germany, pp. 155–73.

Glibert, P. M., Azanza, R., Burford, M., Furuya, K., Abal, E., Al-Azri, A.,
Al-Yamani, F., et al. 2008. Ocean urea fertilization for carbon
credits poses high ecological risks. Mar. Pollut. Bull. 56:1049–56.

Goffart, A., Hecq, J. H. & Legendre, L. 2002. Changes in the
development of the winter-spring phytoplankton bloom in the
Bay of Calvi (northwestern Mediterranean) over the last two
decades: a response to changing climate. Mar. Ecol. Prog. Ser.
236:45–60.

Graneli, E. & Moreira, M. O. 1990. Effects of river water of different
origin on the growth of marine dinoflagellates and diatoms in
laboratory cultures. J. Exp. Mar. Biol. Ecol. 136:89–106.

Graneli, E. & Turner, J. T. [Eds.] 2006. Ecology of Harmful Algae.
Ecological Studies Series 189. Springer Verlag, Heidelberg,
Germany, 403 pp.

Häder, D. P., Worrest, R. C. & Kumar, H. D. 1991. Aquatic eco-
systems. In van der Leun, J. & Tevini, M. [Eds.] Environmental
Effects of Ozone Depletion. United Nations Environment Pro-
gramme, Nairobi, Kenya, pp. 33–40.

Hales, S., Weinstein, P. & Woodward, A. 2001. Ciguatera (fish
poisoning), El Niño and Pacific sea surface temperatures.
Ecosyst. Health 5:20–5.

Hallegraeff, G. M. 1993. A review of harmful algal blooms and their
apparent global increase. Phycologia 32:79–99.

Hallegraeff, G. M. 1998. Concluding remarks on the autecology of
harmful algal blooms. In Anderson, D. M., Cembella, A. D. &
Hallegraeff, G. M. [Eds.] Physiological Ecology of Harmful Algal

Blooms. NATO ASI Series, Vol. G41. Springer Verlag, Heidel-
berg, Germany, pp. 371–8.

Hallegraeff, G., Hosja, W., Knuckey, R. & Wilkinson, C. 2008. Re-
cent range expansion of the red-tide dinoflagellate Noctiluca
scintillans in Australian coastal waters. IOC-UNESCO Harmful
Algae Newsletter 38:10–11.

Hallegraeff, G. M. & Maclean, J. L. 1989. Biology, Epidemiology and
Management of Pyrodinium Red Tides. International Centre for
Living Aquatic Resources Management, Manila, Conf. Proc.
21, 286 pp.

Hallegraeff, G. M., McCausland, M. J. & Brown, R. K. 1995. Early
warning of toxic dinoflagellate blooms of Gymnodinium catena-
tum in southern Tasmanian waters. J. Plankton Res. 17:1163–76.

Hays, G. C., Richardson, A. J. & Robinson, C. 2005. Climate change
and marine plankton. Trends Ecol. Evol. 20:337–44.

Hays, G. C., Warner, A. J., John, A. W. G., Harbour, D. S. &
Holligan, P. M. 1995. Coccolithophores and the continuous
plankton recorder survey. J. Mar. Biol. Assoc. U. K. 75:503–6.

Hayward, T. 1997. Pacific Ocean climate change: atmospheric
forcing, ocean circulation and ecosystem response. Trends Ecol.
Evol. 12:150–4.

Heinze, C. 2004. Simulating oceanic CaCO3 export production in
the greenhouse. Geophys. Res. Lett. 31: L16308, doi: 10.1029/
2004GL020613.

Hoegh-Guldberg, Ø. 1999. Climate change, coral bleaching and the
future of the world’s coral reefs. J. Mar. Freshw. Res. 50:839–66.

Hutchins, D. A., Fu, F. X., Zhang, Y., Warner, M. E., Feng, Y.,
Portune, K., Bernhardt, P. W. & Mulholland, M. R. 2007. CO2
control of Trichodesmium N2 fixation, photosynthesis, growth
rates, and elemental ratios: implications for past, present, and
future ocean biogeochemistry. Limnol. Oceanogr. 52:1293–304.

Iglesias-Rodriguez, M. D., Halloran, P. R., Rickaby, R. E. M., Hall, I.
R., Colmenero-Hidalgo, E., Gittins, J. R., Green, D. R. H., et al.
2008. Phytoplankton calcification in a high-CO2 world. Science
320:336–40.

Imai, I., Yamaguchi, M. & Watanabe, M. 1998. Ecophysiology, life
cycle, and bloom dynamics of Chattonella in the Seto Inland
Sea, Japan. In Anderson, D. M., Cembella, A. D. & Hallegraeff,
G. M. [Eds.] Physiological Ecology of Harmful Algal Blooms. NATO
ASI Series, Vol. G41. Springer Verlag, Heidelberg, Germany,
pp. 95–112.

Integrated Marine Observing System (IMOS). 2009. Available at:
http://imos.org.au/ (last accessed 14 July 2009).

IPCC. 2008. Climate Change 2007 – Impacts, Adaptation and Vulnera-
bility. Working Group II contribution to the Fourth Assessment
Report of the IPCC Intergovernmental Panel on Climate
Change. Available at: http://www.ipcc.ch/ (last accessed 16
January 2010).

Jeffrey, S. W., MacTavish, H. S., Dunlap, W. C., Vesk, M. & Gro-
enewoud, K. 1999. Occurrence of UV A- and UV B-absorbing
compounds in 152 species (206 strains) of marine microalgae.
Mar. Ecol. Prog. Ser. 189:35–51.

Kimura, T., Mizokami, A. & Hashimoto, T. 1973. The red tide that
caused severe damage to the fishery resources in Hiroshima
Bay: outline of its occurrence and environmental conditions.
Bull. Plankton Soc. Jpn. 19:82–96.

Langer, G., Nehrke, G., Probert, I., Ly, J. & Ziveri, P. 2009. Strain-
specific responses of Emiliania huxleyi to changing seawater
carbonic chemistry. Biogeosci. Disc. 6:4361–83.

Lefebvre, K. A., Bargu, S., Kieckhefer, T. & Silver, M. W. 2002. From
sanddabs to blue whales: the pervasiveness of domoic acid.
Toxicon 40:971–7.

Lenton, A., Codron, F., Bopp, L., Metzl, N., Cadule, P., Tagliabue,
A. & Le Sommer, J. 2009. Stratospheric ozone depletion re-
duces ocean carbon uptake and enhances ocean acidification.
Geophys. Res. Lett. 36:L12606. doi: 10.1029/2009GL038227.

Li, W. K. W. 1980. Temperature adaptation in phytoplankton: cel-
lular and photosynthetic characteristics. In Falkowski, P. [Ed.]
Primary Productivity in the Sea. Plenum Press, New York, pp.
259–79.

Lilly, E. L., Kulis, D. M., Gentien, P. & Anderson, D. M. 2002.
Paralytic shellfish poisoning toxins in France linked to a
human-introduced strain of Alexandrium catenella from the

234 G U S T A A F M . H A L L E G R A E F F

western Pacific: evidence from DNA and toxin analysis.
J. Plankton Res. 24:443–52.

Longhurst, A., Sathyendranath, S., Platt, T. & Caverhill, C. 1995. An
estimate of global primary production in the ocean from sa-
tellite radiometer data. J. Plankton Res. 17:1245–71.

Lorius, C., Jouzel, J., Raynaud, D., Hansen, J. & Le Treut, H. 1990.
The ice-core record: climate sensitivity and future greenhouse
warming. Nature 347:139–45.

Maclean, J. L. 1989. Indo-Pacific red tides, 1985–1988. Mar. Pollut.
Bull. 20:304–10.

Margalef, R. 1978. Life-forms of phytoplankton as survival alterna-
tives in an unstable environment. Oceanologia Acta 1:493–509.

McMinn, A. 1988. A Late Pleistocene dinoflagellate assemblage
from Bulahdelah, N.S.W. Proc. Linn. Soc. N. S. W. 109:175–81.

McMinn, A. 1989. Late Pleistocene dinoflagellate cysts from Botany
Bay, New South Wales, Australia. Micropaleontology 35:1–9.

McMinn, A. & Sun, X. 1994. Recent dinoflagellates from the
Chatham Rise, east of New Zealand; a taxonomic study. Paly-
nology 18:41–53.

Merico, A., Tyrrell, T., Brown, C. W., Groom, S. B. & Miller, P. I.
2003. Analysis of satellite imagery for Emiliania huxleyi blooms
in the Bering Sea before 1997. Geophys. Res. Lett. 30:1337. doi:
10.1029/2002GL016648.

Moore, S. K., Mantua, J. M., Hickey, B. & Trainer, V. L. 2008a.
Recent trends in paralytic shellfish toxins in Puget Sound,
relationship to climate, and capacity for prediction of toxic
events. Harmful Algae 8:463–77.

Moore, S. K., Trainer, V. L., Mantua, N. J., Parker, M. S., Laws, E. A.,
Baker, L. C. & Fleming, L. E. 2008b. Impacts of climate vari-
ability and future climate change on harmful algal blooms and
human health. Environ. Health 7(Suppl. 2):S4. doi: 10.1186/
1476-069X-7-S2-S4.

Munhoven, G. 2007. Glacial-interglacial rain ratio changes: impli-
cations for atmospheric CO2 and ocean-sediment interaction.
Deep-Sea Res. Part I Oceanogr. Res. Pap. 54:722–46.

NSTC Joint Subcommittee on Ocean Science and Technology.
2007. Charting the Course for Ocean Science in the United States for
the Next Decade: An Ocean Research Priorities Plan and Implemen-
tation Strategy. Available at: http://ocean.ceq.gov/about/docs/
orppfinal (last accessed 14 July 2009).

Orr, J. C., Fabry, V. J., Aumont, O., Bopp, L., Doney, S. C., Feely, R.
A., Gnanadesikan, A., et al. 2005. Anthropogenic ocean acid-
ification over the twenty-first century and its impact on calci-
fying organisms. Nature 437:681–6.

Park, M. G., Kim, S., Kang, Y. G. & Yih, W. 2006. First successful
culture of the marine dinoflagellate Dinophysis acuminata.
Aquat. Microb. Ecol. 45:101–6.

Peperzak, L. 2005. Future increase in harmful algal blooms in the
North Sea due to climate change. Water Sci. Technol. 51:31–6.

Rahmstorf, S. 2002. Ocean circulation and climate during the past
120,000 years. Nature 419:207–14.

Raven, J., Caldeira, K., Elderfield, H., Hoegh-Guldberg, O., Liss, P.,
Riebesell, U., Shepherd, J., Turley, C. & Watson, A. 2005b.
Ocean Acidification Due to Increasing Atmospheric Carbon Dioxide.
The Royal Society, London, 68 pp.

Raven, J. A., Finkel, Z. V. & Irwin, A. J. 2005a. Picophytoplankton:
bottom-up and top-down controls on ecology and evolution.
Vie Milieu 55:209–15.

Reid, P. C., Johns, D. G., Edwards, M., Starr, M., Poulin, M. &
Snoeijs, P. 2007. A biological consequence of reducing Arctic
ice cover: arrival of the Pacific diatom Neodenticula seminae in
the North Atlantic for the first time in 800,000 years. Glob.
Change Biol. 13:1910–21.

Rhodes, L. L., Haywood, A. J., Ballantine, W. J. & MacKenzie, A. L.
1993. Algal blooms and climate anomalies in North-east New
Zealand, August–December 1992. N. Z. J. Mar. Freshw. Res.
27:419–30.

Richardson, A. J. & Schoeman, D. S. 2004. Climate impact on
plankton ecosystems in the Northeast Atlantic. Science
305:1609–12.

Richardson, A. J., Walne, A. W., John, A. W. G., Jonas, T. D.,
Lindley, J. A., Sims, D. W., Stevens, D. & Witt, M. 2006. Using
continuous plankton recorder data. Prog. Oceanogr. 68:27–74.

Riebesell, U., Schulz, K. G., Bellerby, R. G. J., Botros, M.,
Fritsche, P., Meyerhöfer, M., Neill, C., et al. 2007. Enhanced
biological carbon consumption in a high CO2 ocean. Nature
450:545–8.

Riebesell, U., Zondervan, I., Rost, B., Tortell, P. D., Zeebe, R. E. &
Morel, F. M. M. 2000. Reduced calcification of marine plank-
ton in response to increased atmospheric CO2. Nature
407:364–7.

Rost, B. & Riebesell, U. 2004. Coccolithophores and the biological
pump: responses to environmental changes. In Thierstein, H.
R. & Young, J. R. [Eds.] Coccolithophores: From Molecular Processes
to Global Impact. Springer Verlag, Heidelberg, Germany, pp.
99–125.

Rost, B., Zondervan, I. & Wolf-Gladrow, D. 2008. Sensitivity of
phytoplankton to future changes in ocean carbonate chemis-
try: current knowledge, contradictions and research direc-
tions. Mar. Ecol. Prog. Ser. 373:227–37.

Schippers, P., L}urling, M. & Scheffer, M. 2004. Increase of atmo-
spheric CO2 promotes phytoplankton productivity. Ecol. Lett.
7:446–51.

Shears, N. T. & Ross, P. M. 2009. Blooms of benthic dinoflagellates
of the genus Ostreopsis: an increasing and ecologically impor-
tant phenomenon on temperate reefs in New Zealand and
worldwide. Harmful Algae 8:916–25.

Shelly, K., Heraud, P. & Beardall, J. 2002. Nitrogen limitation in
Dunaliella tertiolecta Butcher (Chlorophyceae) leads to in-
creased susceptibility to damage by ultraviolet-B radiation but
also increased repair capacity. J. Phycol. 38:1–8.

Smayda, T. J. 1990. Novel and nuisance phytoplankton blooms in
the sea: evidence for a global epidemic. In Granéli, E., Sun-
dström, B., Edler, L. & Anderson, D. M. [Eds.] Toxic Marine
Phytoplankton. Elsevier, New York, pp. 29–40.

Smayda, T. J. & Reynolds, C. S. 2001. Community assembly in
marine phytoplankton; application of recent models to
harmful dinoflagellate blooms. J. Plankton Res. 23:447–61.

Stachowicz, J. J., Terwin, J. R., Whitlatch, R. B. & Osman, R. W.
2002. Linking climate change and biological invasions: ocean
warming facilitates non-indigenous species invasion. Proc. Natl.
Acad. Sci. U. S. A. 99:15497–500.

Sunda, W. G., Graneli, E. & Gobler, C. J. 2006. Positive feedback
and the development and persistence of ecosystem disruptive
algal blooms. J. Phycol. 42:963–74.

Tester, P. A., Geesey, M. E. & Vukovich, F. M. 1993. Gymnodinium
breve and global warming: what are the possibilities? In Smayda,
T. J. & Shimizu, Y. [Eds.] Toxic phytoplankton blooms in the
sea. Dev. Mar. Biol. 3:67–72.

Tester, P. A., Stumpf, R. P., Vukovich, F. M., Folwer, P. K. & Turner,
J. T. 1991. An expatriate red tide bloom: transport, distribu-
tion, and persistence. Limnol. Oceanogr. 36:1053–61.

Thorsen, T., Dale, B. & Nordberg, K. 1995. Blooms of the toxic
dinoflagellate Gymnodinium catenatum as evidence of climatic
fluctuations in the Late Holocene of southwestern Scandina-
via. The Holocene 5:435–46.

Tortell, P. D., Giocoma, R. D., Sigman, D. M. & Morel, F. M. M.
2002. CO2 effects on taxonomic composition and nutrient
utilization in an equatorial Pacific phytoplankton assemblage.
Mar. Ecol. Prog. Ser. 236:37–43.

Turner, J. T. & Graneli, E. 2006. ‘‘Top-down’’ predation control on
marine harmful algae. In Graneli, E. & Turner, J. T. [Eds.]
Ecology of Harmful Algae. Ecological Studies Series 189.
Springer, Verlag, Heidelberg, Germany, pp. 355–66.

Van Dolah, F. V. 2000. Marine algal toxins: origins, health effects,
and their increased occurrence. Environ. Health Perspect.
108(Suppl. 1):133–41.

Walsh, J. J. & Steidinger, K. A. 2001. Saharan dust and Florida red
tides: the cyanophyte connection. J. Geophys. Res. 106:11597–
612.

Woods, J. & Barkmann, W. 1993. The plankton multiplier –
positive feedback in the greenhouse. J. Plankton Res. 15:
1053–74.

Yin, K. D., Harrison, P. J., Chen, J., Huang, W. & Qian, P. Y. 1999.
Red tides during spring in 1998 in Hong Kong: is El Niño
responsible? Mar. Ecol. Prog. Ser. 187:289–94.

C L I M A T E C H A N G E A N D A L G A L B L O O M S 235

REVIEW

TESTING THE EFFECTS OF OCEAN ACIDIFICATION ON ALGAL METABOLISM:
CONSIDERATIONS FOR EXPERIMENTAL DESIGNS

1

Catriona L. Hurd,2 Christopher D. Hepburn

Department of Botany, University of Otago, PO Box 56, Dunedin 9054, New Zealand

Kim I. Currie

National Institute for Water and Atmospheric Research Ltd., Centre of Excellence for Chemical and Physical Oceanography,

Department of Chemistry, University of Otago, PO Box 56, Dunedin 9054, New Zealand

John A. Raven

Division of Plant Sciences, Scottish Crop Research Institute, University of Dundee at SCRI, Invergowrie, Dundee DD2 5DA, UK

and Keith A. Hunter

Department of Chemistry, University of Otago, PO Box 56, Dunedin 9054, New Zealand

Ocean acidification describes changes in the car-
bonate chemistry of the ocean due to the increa-
sed absorption of anthropogenically released CO2

.

Experiments to elucidate the biological effects of
ocean acidification on algae are not straightforward
because when pH is altered, the carbon speciation
in seawater is altered, which has implications for
photosynthesis and, for calcifying algae, calcifica-
tion. Furthermore, photosynthesis, respiration, and
calcification will themselves alter the pH of the sea-
water medium. In this review, algal physiologist

s

and seawater carbonate chemists combine their
knowledge to provide the fundamental information
on carbon physiology and seawater carbonate chem-
istry required to comprehend the complexities of
how ocean acidification might affect algae metabo-
lism. A wide range in responses of algae to ocean
acidification has been observed, which may be
explained by differences in algal physiology, time-
scales of the responses measured, study duration,
and the method employed to alter pH. Two meth-
ods have been widely used in a range of experimen-
tal systems: CO2 bubbling and HCl ⁄ NaOH
additions. These methods affect the speciation of
carbonate ions in the culture medium differently; we
discuss how this could influence the biological
responses of algae and suggest a third method based
on HCl ⁄ NaHCO3 additions. We then discuss eight
key points that should be considered prior to setting
up experiments, including which method of manipu-
lating pH to choose, monitoring during experiments,
techniques for adding acidified seawater, biological

side effects, and other environmental factors. Finally,
we consider incubation timescales and prior condi-
tioning of algae in terms of regulation, acclimation,
and adaptation to ocean acidification.

Key index words: algae; bicarbonate; calcium car-
bonate; carbon; carbon dioxide; climate change;
ocean acidification; phytoplankton; seawater car-
bonate system; seaweed

Abbreviations: AT, total alkalinity; CA, carbonic
anhydrase; CCM, carbon-concentrating mecha-
nism; CT, total inorganic carbon; pCO2, partial
pressure of CO2(g

)

The term ‘‘ocean acidification’’ describes changes
in the carbonate chemistry of the ocean due to
increased CO2 absorption since the Industrial Revo-
lution (The Royal Society 2005, Doney et al. 2009

).

Phytoplankton and macroalgae have key ecological
roles as primary producers in coastal and open
oceans, supplying fixed carbon to the entire marine
food web, recycling nutrients, and modifying global
climate (Smith 1981, Duggins et al. 1989, Field et al.
1998, Gattuso et al. 1998a, Zondervan 2007). Addi-
tionally, calcareous algae (e.g., planktonic cocco-
lithophores and benthic calcifying macroalgae) are
a major source of marine carbonates and sediments
(Gattuso et al. 1998b, Feely et al. 2004, Balch et al.
2007), and the coralline macroalgae fulfill impor-
tant ecological processes, including reef building
(Adey 1998, Chisholm 2003), and are the preferred
sites for settlement of invertebrate larvae (Roberts
2001).

A major focus of research into ocean acidification
has been the effects on calcifying organisms

1Received 28 November 2008. Accepted 6 May 2009.
2Author for correspondence: e-mail catriona.hurd@botany.otago.

ac.nz.

J. Phycol. 45,

1236

–1251 (2009)
� 2009 Phycological Society of Americ

a

DOI: 10.1111/j.1529-8817.2009.00768.x

1236

(animals and algae). Elevated seawater CO2 concen-
trations will lower carbonate saturation states, which
in turn may reduce the ability of calcifiers to main-
tain existing, and build new, carbonate skeletons
(Bijma et al. 1999, Riebesell et al. 2000, Orr et al.
2005, Shirayama and Thornton 2005, De’ath et al.
2009). These effects will occur in high latitudes first;
for example, in Southern Ocean surface waters,
undersaturation of aragonite is predicted between
2030 and 2038 (McNeil and Matear 2008). Such
reduced ability to calcify may decrease their compet-
itive fitness (Kuffner et al. 2007, McNeil and Matear
2008). In addition to influencing calcification, the
changed speciation of dissolved inorganic carbon in
seawater and decreased pH via the carbonate buffer
system, and the differing abilities of algae to utilize
CO2 and HCO3

), ocean acidification has the poten-
tial to affect the metabolism and growth rates of all
algae, both noncalcareous and calcareou

s.

Ocean acidification is an emerging field of
research. Experiments to elucidate the biological
effects of increased CO2 on marine organisms are
not straightforward because when pH is altered, the
carbon speciation in seawater is modified, which has
strong implications for photosynthesis, respiration,
and calcification. Furthermore, these same three
metabolic processes themselves alter the pH of sea-
water medium surrounding the algae. Therefore, an
understanding of both seawater carbonate chemistry
and physiological processes related to carbon metab-
olism and calcification is required to design experi-
ments without artifacts, which can be carefully
replicated and test the impacts of increased CO

2

(i.e., lowered pH) on algae. This review is in two
sections. In the first section, we highlight key
aspects of seawater carbonate chemistry, algal car-
bon acquisition, and calcification and consider the
wide range of biological responses by different algae
to ocean acidification. On the basis of this appraisal,
we consider reasons for the observed broad range
of physiological responses, which include physiology
of different species, timescales of studies, and tech-
niques used to modify seawater pH (CO2 bubbling
vs. HCl ⁄ NaOH additions). In section two, we focus
on how bubbling with CO2 and adding HCl ⁄ NaOH
each modifies the carbonate chemistry of seawater
during incubation studies, discuss the probable bio-
logical responses of algae to each technique, and
recommend a series of steps that should be consid-
ered when designing experiments to test the effects
of ocean acidification on algae.

Carbon chemistry and biological responses to its manip-
ulation. Seawater carbonate chemistry: Since the Indus-
trial Revolution, levels of CO2 in the atmosphere
have increased at rates 100-fold greater than prein-
dustrial times and have caused a rise in atmospheric
CO2 levels from 280 to 384 ppmv (Solomon et al.
2007). CO2 in the atmosphere is increasing at rates
faster than that predicted as a ‘‘worst-case scenario’’

by the International Panel for Climate Change
(IPCC) in 2000 (Raupach et al. 2007). The world’s
oceans have absorbed up to 50% of anthropogeni-
cally derived CO2, and modeling studies suggest a
0.1 unit decline in surface seawater pH since 1750
(Caldiera and Wickett 2003).

In seawater, free CO2(aq) is in equilibrium with a
small concentration of the carbonic acid species
H2CO3, but it is conventional to regard both species
as stoichiometrically equivalent with respect to sub-
sequent acid-base reactions and denote this combi-
nation by a hypothetical species H2CO3

* where
[H2CO3

*] = [H2CO3] + [CO2(aq)]. The effects of
ocean acidification cannot be described in a simple
way using just the parameter pH, and it is necessary
to consider the effect of CO2 uptake on the entire
CO2 equilibrium system. CO2 in the gas phase equil-
ibrates with H2CO3

* in seawater through the well-
known Henry’s law equilibrium:

CO2ðgÞ$ H2CO3� KH ¼
½H2CO3��

pCO2
ð1Þ

where pCO2 is the partial pressure of CO2(g) and
KH, the Henry’s law equilibrium constant, is a func-
tion of temperature (T) and salinity (S). This rela-
tionship means that at a given T and S, pCO2 and
[H2CO3

*] are linearly related to each other. It is
most common to use pCO2 as a parameter because
this allows a simple comparison with the actual
atmospheric CO2 partial pressure when air and
water are not in equilibrium.

The acid dissociation reactions of H2CO3
* are as

follows:

H2CO3
�$HCO3�þHþ K1¼

½HCO3��½Hþ�
½H2CO3��

ð2Þ

HCO3
�$CO32�þHþ K2¼

½CO32��½Hþ�
½HCO3��

ð3Þ

where HCO3
) and CO3

2) are the bicarbonate and car-
bonate ions, respectively, and K1 and K2 are the first
and second acid dissociation constants of H2CO3

*.
Equations (2) and (3) show that the concentrations
of the three CO2 species and that of H

+ are inextrica-
bly linked, meaning that it is physically impossible to
vary systematically any one of these while at the same
time holding all of the others constant. Equation (1)
shows that this relationship also extends to pCO2.
This fact complicates understanding the underlying
chemistry affecting ocean acidification.

Concentrations of the individual CO2 species in
seawater cannot be directly measured. Instead,
changes in the speciation of the CO2 system in sea-
water are normally described, and measured, using
the following two parameters (Mackenzie and
Lerman 2006): (i) Total dissolved CO2 (usually
symbolized as CT for the total inorganic carbon in
solution or DIC for dissolved inorganic carbon),

O C E A N A C I D I F I C A T I O N A N D A L G A L M E T A B O L I S M 1237

which is the stoichiometric sum of all dissolved inor-
ganic carbon species

C T ¼ ½H2CO3��þ ½HCO3��þ ½CO32�� ð4Þ

(ii) Total alkalinity, AT, which is the total concentra-
tion of titratable weak bases in seawater relative to
the reference proton condition comprising pure
CO2 in seawater

AT¼½HCO3��þ2½CO32��þ½OH���½Hþ�þð…Þ ð5Þ
where (…) represents various minor acid-base spe-
cies, such as borate ion. Both these parameters have
the advantage of being independent of changes in
temperature and pressure and are conserved during
the mixing of different seawater masses. Useful soft-
ware programs for calculation of CO2 speciation in
seawater have been presented by Lewis and Wallace
(1998) and Hunter (2007).

For calcification, the removal of CO3
2) ions by

precipitation of calcium carbonate (CaCO3) causes
HCO3

) ions to dissociate to restore the CO3
2) ions

lost. The H+ ion released by this dissociation gener-
ates additional H2CO3

* by combining with a second
HCO3

) ion (Frankignoulle and Canon 1994). The
overall stoichiometric change is therefore:

Ca2þ þ 2HCO3� ! CaCO3ðsÞþ H2CO3� ð6Þ

At today’s pH (�8.07), 91% of CT is as bicarbon-
ate ions (2,200 lM), 1% as H2CO3

* (14 lM), and
8% as CO3

2). The predicted decrease to pH 7.65 by
2100 will result in a 300% increase in H2CO3

* con-
centration, a 9% increase in that of HCO3

), and a
56% decrease in that of CO3

2) (from table 1 of The
Royal Society 2005). These changes will affect the
ability of algae to acquire carbon and ⁄ or produce
and maintain calcium carbonate structures.

Physiological basis for algal carbon acquisition and
calcification: Most marine algae can acquire the
CO2 required as a substrate for RUBISCO via active
uptake from seawater of CO2 and ⁄ or bicarbonate;
the active transport of either of these species, or in
some cases of protons, constitutes a carbon-concen-
trating mechanism (CCM; Giordano et al. 2005).
The photosynthetic rates of algae that have CCMs
are not generally carbon limited under most envi-
ronmental conditions (Giordano et al. 2005). Some
bicarbonate-using algae convert HCO3

) to CO2
using extracellular carbonic anhydrase (CA); the
CO2 then enters the cell by active transport or by
diffusion (if there are zones of surface acidification
where the steady-state CO2 concentration exceeds
that in the medium). Other bicarbonate-using algae
with CCMs actively take up the HCO3

) ion across
the cell membrane, and CA acts intracellularl

y.

Some algae (e.g., some dinoflagellates) have little or
no capacity to use bicarbonate, and their CCM relies
on active CO2 uptake (Dason et al. 2004). CA syn-
thesis and CCM activities in eukaryotes are con-

trolled by the concentration of H2CO3
*, in the few

cases examined (Giordano et al. 2005).
Energy and nutrients are required to operate

active transport and to make the CCMs (generally
including CAs), whereas the alternative of diffusive
H2CO3

* has energy and nutrient costs of operating
photorespiration and making the relevant enzymes
and additional RUBISCO (Raven et al. 2000). Some
algae adapted to low light levels lack CCMs and rely
on diffusive H2CO3

* entry (e.g., the red seaweed
Lomentaria, Kübler et al. 1999). Under low irradianc-
es, energy limitation outweighs limitation by CO2,
and the use of diffusive H2CO3

* entry has energetic
advantages (see Raven et al. 2000, 2005). Growth at
low irradiances cannot explain all cases of the
absence of detectable CCMs; for example, some
strains of coccolithophores rely on diffusive uptake
of H2CO3

* and cannot utilize HCO3
) or carry out

active CO2 transport (Nimer and Merrett 1993).
Algae that rely on H2CO3

* diffusion alone are gen-
erally carbon-limited under today’s seawater concen-
trations (Kübler et al. 1999).

Calcification is the biogenic formation of calciu

m

carbonate (Borowitzka 1987). The most common
forms of calcium carbonate (CaCO3) synthesized by
algae are calcite, aragonite, or high-magnesium cal-
cite (Adey 1998). High magnesium calcite is the
most soluble form of these three (Chave et al.
1962), and therefore algae with high-magnesium
calcite are most susceptible to predicted decreases
in pH due to ocean acidification (The Royal Society
2005). Algae have a range of mechanisms for calcify-
ing. For example, coccolithophores produce calcite
coccoliths intracellularly and extrude them to the
cell’s surface, whereas in the tropical green seaweed
Halimeda, aragonite mineralizes in the intercellular
space between tightly appressed utricles, and red
coralline seaweeds deposit high-magnesium calcite
into their cell walls (Borowitzka 1987).

Algae themselves modify the pH of seawate

r.

When algae photosynthesize, the removal of CO2 in
assimilation by RUBISCO usually occurs faster than
CO2 can be resupplied from the atmosphere or dee-
per water, so that there is a reequilibration among
the inorganic species that yields a decrease in
HCO3

) and an increase in CO3
2) and pH. In nat-

ure, this results in significant pH increases most
especially in isolated habitats like high-intertidal
rock pools (Midelboe and Hansen 2007) but also in
coastal waters (Hinga 2002). Indeed, such effects
are the principle underpinning pH-drift experi-
ments where a rapid increase in pH as a result of
photosynthesis is used to elucidate mechanisms of
carbon acquisition by seaweeds and microalgae
(Maberly 1990, Chen et al. 2006). Calcification and
respiration alter the seawater carbonate system in
ways that decrease the pH of the culture medium.
Respiration results in CO2 being released into the
surrounding medium, and when algae are in the
dark, the pH of the culture medium will decline.

1238 C A T R I O N A L . H U R D E T A L .

T
a

b
l
e

1
.

N
o

n
ca

lc

a
re

o
u

s

a
lg

a
e

.

E
x

a
m

p
le

s
o

f

e
x

p
e
ri

m

e
n

t

s
co

n
d

u
ct

e
d

to
te

st

th

e

e
ff

e
ct

s
o

f
C

O
2

c

o
n

ce
n

tr

a
ti

o
n

in
se

a
w

a
te

r
o

n
a

lg

a
l

g
ro

w
th

a
n

d
⁄o

r
m

e
ta

b
o

li
sm

.
U

n
it

s
fr

o
m

th
e

o
ri

g
in

a
l

p
a
p

e

rs

a
re

r

e
p

o
rt

e
d

.
L

😀
,

li
g

h
t:

d

a
rk

.

S
p

e
ci

e
s

⁄s
ys

t

e
m

M
e
th

o
d

s
T

im
e

fr
a
m

e
p

H
⁄p

C
O

2

tr
e

a
tm

e
n

ts
R

e
p

li
ca

ti
o

n
R

e
su

lt
s

A
u

t

h
o

rs

P
h

y

t
o

p
la

n
k

to
n

A
st

er
io

n
el

la
gl

a
ci

a
li

s,
T

h
a
la

ss
io

si
ra

p
u

n
ct

ig
er

a
,

C
os

ci
n

od
is

c

u
s

w
a
il

es
ii

,
P

h
a
eo

d
a
ct

yl
u

m
tr

ic
or

n
u

t

u
m

,
S
cr

i

p
p

si

el
la

tr
oc

h
oi

d
ea

H
C

l⁄
N

a
O

H
,

d
il

u
te

b

a
tc

h
cu

lt
u

re
s,

1
5

0
l

m
o

l
p

h
o
to
n

s
Æ

m
)

2
Æ

s

)
1

,

1
6

:8
o

r
2

4
:0

L
😀

,
f
⁄2

n
u

tr
ie

n
ts

.

C
o

m
p

le
te

a
cc

li
m

a
ti
o
n

o
f

cu
lt

u
re

s
to

g
ro
w
th

co
n

d
it

io
n

s
li

k
e
ly

.

S
ix

C
O
2
co
n
ce
n

tr
a
ti

o
n

s
ra

n
g

in
g

fr
o

m
1

to
3

8
l

m
o

l
Æ

k
g

)
1
N
o

t
cl

e
a
r

N
o
e
ff
e
ct
o
f

tr

e
a
tm

e
n
t
o
n
g

r

o
w

th
ra

te
o

n
a

c

e
ll

in
cr

e
a
se

o
r

o
rg

a
n

ic
ca

rb
o

n
b

a
si

s.

B
u

rk
h

a
rd

t

e
t

a
l.

(1
9

9
9

)
P
h

yt
o

p
la
n
k
to
n

a
ss

e
m

b
la

g
e

s,

e
q

u
a
to

ri
a

l

P
a
ci

fi
c

o
ff

P
e
ru

C
O

2
b

u
b

b
li

n
g

,
se

m
ic

o
n

ti
n

u
o

u
s

b
a
tc

h
cu
lt
u

re
,

3
0

%
su

rf

a
ce

ir
ra

d
ia

n
ce

,
su

rf
a
ce

se
a
w

a
te

r
te

m
p

e
ra

tu
re

,
n

a
tu

ra
l

n
u
tr
ie
n
ts
.
E
x
p
e
ri

m
e
n

ts
la

st
e
d

1
1

d
.

p
C

O
2
1
5

0
(j

u
st

le
ss

th
a
n

l

a
st

g
la

ci
a
l

m
a
x

i

m
u

m
)
a
n

d
7

5
0

p
p
m

n
=

3
N

o
si

g
n

ifi
ca

n
t

e
ff
e
ct

ch
a
n

g
e
d

C
O

2
o

n
p

ri
m

a
ry

p
ro

d
u

ct
io

n
(1

4
C

)
o

f
th

e
a
ss

e
m
b
la

g
e
.

N
a
n

o
fl

a
g

e
ll
a
te

s
(u

n
ic

e
ll

u
la

r
P

h
a
eo

cy
st

is
)

le
ss

si
g

n
ifi

ca
n

t
re

la
ti

ve
to

d
ia

to
m

s
in

h
ig

h
C

O

2
.

T
o

rt
e
ll

e
t
a
l.

(

2
0

0
2

)

E
m

il
ia

n
ia

h
u

xl
ey

i
(n

o
n

c

a
lc

if
yi

n
g

s

t
ra

in
)

C
O
2
b
u
b
b
li
n
g

,
ch

e
m

o
st

a
t

a
n

d
b

a
tc
h
cu
lt
u
re
s,

8
0

o
r
5
0
0
l
m
o
l
p
h
o
to
n
s
Æ
m
)
2
Æ

s)
1

ir
ra
d
ia
n
ce

tr
e
a
tm

e
n

ts
,

1
4

:

1
0

L
😀

,
1

8
�C

,
a
ls

o
m
a
n

ip
u

la
te

d
n

u
tr

ie
n

ts
.

1
p

lu
s

3
w

e
e
k

s
a
cc

li
m
a
ti
o
n

to
e
x

p
e
ri
m
e
n

ta
l

tr
e
a
tm
e
n
ts
.
p
C
O
2

=
3

6
0

a
n

d
2

,

0
0

0
p

p
m

n
=

3
In

cr
e
a
se

d
g

ro
w

th
a
t

h
ig

h

e
r

p
C
O
2
tr
e
a
tm
e
n
t
ra

te
p

e
r
o
rg
a
n

ic
m

a
tt

e
r,

b
u

t
n

o
in

cr
e
a
se
o
n
a
ce

ll
n

u
m

b
e
r

b
a
si

s.

L
e
o

n
a
rd

o
s

a
n

d
G

e
id

e
r

(2
0

0
5

)

P
ro

ch
lo

ro
co

cc
u

s
C

C
M

P
1

9
8

6
,

S
yn

ec
h
oc

cu
s

sp
.

C
C

M
P

1
3

3
4

C
O
2
b
u
b
b
li
n
g
,
se
m
ic
o
n
ti
n
u
o
u
s
cu
lt
u
re

,
4

5
l

m
o
l
Æ
m
)
2
Æ

s)

1
,

1
2

:1
2

L
😀

,
2

0
�C

a
n
d
2

4
�C

,
o

li
g

o
tr

o
p

h
ic

n
a
tu

ra
l
se
a
w
a
te

r
w

it
h

A
Q

U
IL

n
u
tr
ie
n
ts
.

A
t

le
a
st

2
w

e
e
k

s
(a

t
le

a
st

7
g

e
n
e
ra
ti
o

n
s)

a
cc
li
m
a
ti
o
n
.
p
C
O
2

3
8

0
a
n

d
7
5
0
p
p

m
n

=
3
N
o
si
g
n
ifi
ca
n

t
e
ff

e
ct
o
n
g
ro
w
th

ra
te

o
f

e
it

h
e
r

o
rg
a
n

is
m

o
f
in
cr
e
a
se

d
C

O
2
a
t
a
g

i

v

e
n

te
m

p
e
ra

tu
re
,
o
r
o

n
o

rg
a
n

ic
C

p
e
r

ce
ll

in
P

ro
ch

lo
ro

co
cc

u
s.

In
cr

e
a
se

d
o

rg
a
n
ic
C
p
e
r
ce
ll

w
it

h
in

cr
e
a
se
d
C
O
2

in
S
yn

ec
h
oc

cu
s.

F
u

e
t
a
l.
(2
0

0
7

)

O C E A N A C I D I F I C A T I O N A N D A L G A L M E T A B O L I S M 1239

T
a
b
l
e
1
.
C
o

n
ti

n
u
e
d
.
S
p
e
ci
e
s
⁄s
ys
te
m
M
e
th
o
d
s
T
im
e
fr
a
m
e
p
H
⁄p
C
O
2
tr
e
a
tm
e
n
ts
R
e
p
li
ca
ti
o
n
R
e
su
lt
s
A
u

th
o

rs
P
h
yt
o
p
la
n
k
to
n
a
ss
e
m
b
la
g
e
s,

R
o

ss
S

e
a
,

A
n

ta
rc

ti
ca

C
O
2
b
u
b
b
li
n
g

,
sh

ip
b

o
a
rd

,
se
m
ic
o
n
ti
n
u
o
u
s
b
a
tc
h
cu
lt
u
re
,
3
0

%
a
m

b
ie

n
t
ir
ra
d
ia
n
ce

,
0

�C
,

n
a
tu
ra
l

le
ve

ls
o

f
m

a
cr

o
n
u
tr
ie
n
ts
,

1
n

M
F

e
C

l 3
.

E
x
p
e
ri
m
e
n
ts
la
st
e
d
1
0
o
r

1
8

d
.
p
C
O
2

=
1

0
0

,
3

8
0

,
a
n

d
8

0
0
p
p
m
n
=
3
N
o

v.
–
D

e
c.

e
x
p
e
ri
m
e
n

t:
1
4
C

u
p

ta
k

e
(2

4
h

)
a
n

d
g
ro
w
th
ra

te
s

h
ig
h
e
r

in
3

8
0
a
n
d
8
0
0
tr
e
a
tm
e
n

ts
co

m
p
a
re

d
to

1
0
0
p
p
m

.
Ja

n
.

e
x
p
e
ri
m
e
n
t:
1
4
C
u
p
ta
k

e
in

cr
e
a
se
s
li

n
e

a
rl

y
w

it
h
in
cr

e
a
si

n
g
p
C

O
2
,

a
n
d
g
ro
w
th
ra

te
h

i

g
h

e
r

in
8

0
0
p
p

m
co

m
p
a
re
d
to
1
0
0
a
n

d
3

8
0
p
p

m
.

T
o
rt
e
ll
e
t
a
l.
(2
0

0
8

)

M
a
c
ro

a
lg
a
e

L
om

e

n
ta

ri
a

a
rt

ic
u

la
ta

B
u

b
b

li
n

g
w

it
h
C
O

2
⁄O

2
⁄N

2
m

ix
,

0
.0

6
L

cu
lt
u
re

ve
ss

e
l,

4
0

–
5

0
l
m
o
l
p
h
o
to
n
s
Æ
m
)
2
Æ

s)
1
,

1
2
:1
2
L
😀
,
1
0
�C
,
P

ro
va

s

o
li

e
n

ri
ch

e
d
se
a
w
a
te
r.

1
-w

e
e
k
a
cc
li
m
a
ti
o
n
,
3

-w
e
e
k

e
x
p
e
ri
m
e
n

t.

p

C

O
2

=
2

3
5

,
3
5
0

,
7

0
0
a
n

d
1

,

7
5

0
p
p
m
n
=

4
5

2
%

in
cr
e
a
se
in
g
ro
w
th
ra

te
.

K
ü

b
le

r
e
t

a
l.
(1
9
9
9
)

P
or

p
h

yr

a
li

n
ea

ri
s

In
je

ct
io
n
o
f
C
O
2

g

a
s

u
si

n
g

p
H

co
n

tr
o

ll
e
r

a
n

d
T

A
P

S
b

u
ff

e
r

fo
r

h
ig

h
p

H
,

0
.5

L
cu

lt
u

re
ve

ss
e
l,

7
0

l
m

o
l

p
h

o
to

n
Æ

m
)
2
Æ
s)
1
,
1
0

:1
4

L
😀
,
1

5
�C

,
P
ro
va
so
li

E
n

ri
ch
m
e
n

t
M

e
d

ia
p

lu
s

M
e
ta

l
M

ix
II

.
1
-w
e
e
k
a
cc
li
m
a
ti
o
n
,
2
-w
e
e
k
e
x
p
e
ri
m
e
n
t.
p

H
=

6
.0

,
6

.5
,

7
.0

,
8

.7
N

o
t

cl
e
a
r

N
o
cl
e
a
r

d
if

fe
re

n
ce

s
w

it
h

in
re

a
li

st
ic

ra
n

g
e

(7
.5

–
8

).

Is
ra

e
l

e
t
a
l.
(1
9
9
9
)
C
o

n
st

a
n

t
a
e
ra

ti
o

n
w

it
h
C
O

2
⁄a

ir
m

ix
,
4
0
L
cu
lt
u

re
ta

n
k

s,
o

th
e
r

co
n
d
it
io
n

s
n

o
t

st
a
te

d
.
3
w
e
e
k
s
C
o
n
tr
o

l
=

3
6

0
m

g
Æ

L
)

1

T
re

a
tm
e
n

t
=

7
5
0
m
g
Æ
L
)
1
n
=
3
5

0
%

re
d

u
ct
io
n
in
g
ro
w
th
a
t
h
ig
h
C
O
2
.

1240 C A T R I O N A L . H U R D E T A L .

T
a
b
l
e
1
.
C
o
n
ti
n
u
e
d
.
S
p
e
ci
e
s
⁄s
ys
te
m
M
e
th
o
d
s
T
im
e
fr
a
m
e
p
H
⁄p
C
O
2
tr
e
a
tm
e
n
ts
R
e
p
li
ca
ti
o
n
R
e
su
lt
s
A
u
th
o
rs
1
4

g
re

e
n

,
re

d
,

a
n
d
b
ro
w

n
m

a
cr

o
a
lg

a
e
C
o
n
st
a
n
t
a
e
ra
ti
o
n
w
it
h
C
O
2
⁄a
ir
m
ix
,
4
0
L
cu
lt
u
re
ta
n
k

s,
0

.2
m

M
p

h
o

sp
h

a
te
,
2

m
M

a
m
m
o

n
iu

m
.

4

–
1

5
w

e
e
k
s
C
o
n
tr
o
l
=
3
6
0
l
m
o
l
Æ
m
o

l)
1

T
re
a
tm
e
n
t
=
7
5
0
l
m
o
l
Æ
m
o
l)
1
N
o
t
cl
e
a
r
N
o
cl
e
a
r
e
ff
e
ct
o
f

e
le

va
te

d
C
O
2
o
n
g
ro
w
th
o
r
p
h
o
to

sy
n

th

e
si

s.
L

o
w
p
H

m
a
y

h
a
ve

d
a
m

a
g
e
d

so
m

e
sp

e
ci

e
s.

Is
ra
e
l
a
n

d
H

o
p

h
y

(2
0
0
2
)

P
u

re
C

O
2

in
je

ct
e
d

a
t
th
e

b
a
se

m

e
sh

b
a
g

s
co
n
ta

i

n
in

g
se

a
w

e
e
d

s
in

si
tu

,
a
ve

ra
g

e
1

,

4
9

0
l
m
o
l
p
h
o
to
n
s
Æ
m
)
2
Æ
s)
1
,

2
8

�C
,
a
m
b
ie
n
t
n
u
tr
ie
n
ts
.

5
d

C
o

n
tr

o
l
p
H

=
8

.1
8

–
8

.1
4

T
re
a
tm
e
n

t
p

H
=

7
.4

–
7

.6

3

N
o
t
cl
e
a
r
N
o
cl
e
a
r
e
ff
e
ct
s
o

f
e
le

va
te
d
C
O
2
.

S
a
cc

h

a
ri

n
a
la
ti

ss
im

a
a
n

d
N

er
eo

cy
st

is
lu

et
ke

a
n
a
C
o
n
st
a
n
t
a
e
ra
ti
o
n
w
it
h
C
O
2
⁄a
ir
m

ix
(3

,0
0

0
p
p
m

)
d

ir
e
ct

ly
in

o
u

td
o

o
r
7
5

0
L

ta
n

k
s,

m
a
x

.
2

,2
0

0
l
m
o
l
p
h
o
to
n
s
Æ
m
)
2
Æ
s)
1
,
1
1

�C
–
1

5
�C

,
a
m

b
ie
n
t
n
u
tr
ie
n
ts
.

5
5

d
C
o
n
tr
o
l
p

H
�

8
.0

2
–
8

.1
6

,
tr

e
a
tm
e
n

t
�

7
.1

7
–
7

.9
9

a
t

m
id

d
a
y.

p
H

co
u

ld
b

e
a
s

m
u

ch
a
s
�

0
.8

u
n

it
s

lo
w

e
r
a
t

d
a
w

n
.

P
se

u
d

o
-r

e
p
li
ca

te
d

3
·

in
cr
e
a
se
d
g
ro
w

th
in

S
.

la
tt

is
m
a
,

4
·

d

e
cr

e
a
se

in
N

.
lu

et
ke
a
n

a
.

S
w

a
n

so
n

a
n

d
F

o
x

(2
0
0
7
)

O C E A N A C I D I F I C A T I O N A N D A L G A L M E T A B O L I S M 1241

Calcification also results in CO2 production (see
eq. 6).

Biological responses of algae to pH manipulation: The
goal of most experiments investigating ocean acidifi-
cation is to increase the concentration of H2CO3

*

and decrease the pH in seawater to mimic the
increase in H2CO3

* predicted to occur as the ocean
takes up anthropogenic carbon and examine the
effects (physiological, ecological, biogeochemical)
of such manipulations on algae. Two techniques
have been used to manipulate seawater pH in the
majority of biological perturbation experiments:
CO2 bubbling and HCl ⁄ NaOH additions. These
have been used in a variety of experimental setups
over various incubation timescales (Tables 1 and 2).

Across the range of calcareous and noncalcareous
algae tested, there are no clear patterns regarding
the responses of primary production, growth, or cal-
cification rates to ocean acidification (Tables 1 and
2). The growth rate of some species is unchanged
by altered CO2 treatments—for example, Thalassios-
ira pseudonana (Pruder and Bolton 1980), four dia-
toms and one dinoflagellate (Burkhardt et al.
1999), the coccolithophores Calcidiscus leptoporus
and Coccolithus pelagica (Langer et al. 2006), and
Emiliania huxleyi (Feng et al. 2008). For T. pseudo-
nana, the lack of change in growth rate following
CO2 treatment is consistent with inorganic carbon
concentrations being saturating for growth (Clark
and Flynn 2000). Species that responded to
CO2 ⁄ pH treatments include E. huxleyi, in which
increased H2CO3

* concentrations resulted in
increased organic carbon content per cell, but no
increase in the number of cells (Leonardos and
Geider 2005, Iglesias-Rodriguez et al. 2008a).
Growth rates of Antarctic phytoplankton assem-
blages were also affected by pCO2, but the response
to treatments varied depending on the time of year
when the experiment was conducted (Tortell et al.
2008). The marine diazotrophic cyanobacterium
Trichodesmium also demonstrated significant
increases in the rate of carbon assimilation (and of
diazotrophic nitrogen assimilation) with substantial
increases in CO2 concentration, in each of three
studies (Barcelos e Ramos et al. 2007, Hutchins
et al. 2007, Levitan et al. 2007). Similar results were
reported for the unicellular marine diazotrophic
cyanobacterium Crocosphaera under iron-sufficient,
but not iron-limiting conditions (Fu et al. 2008).

For calcareous E. huxleyi and Ca. leptoporus, there
was a decrease in rates of calcification and ⁄ or mal-
formed coccoliths at pCO2 values greater (or lower)
than the present day; however, for Co. pelagica, there
was no effect of pCO2 treatment on lith formation
(Riebesell et al. 2000, Langer et al. 2006). Interest-
ingly, when nanofossil records from cores from the
last glacial maximum (�18,000 years ago, atmo-
spheric CO2 180–200 lmol Æ mol

)1 total gas) were
examined, there was no evidence of malformed or
incomplete liths for Ca. leptoporus or Co. pelagica;

Langer et al. (2006) suggest adaptation (see below
for definition) by these species to the pCO2 environ-
ment they inhabit. Iglesias-Rodriguez et al. (2008a)
found no decrease in calcification or lith malforma-
tion in E. huxleyi grown at pCO2 higher than the
present day.

Relatively few studies have determined the likely
impact of elevated CO2 on calcareous and noncal-
careous macroalgae. Earlier works (unrelated to
ocean acidification) used manipulations of seawater
pH and carbon chemistry to unravel mechanisms of
carbon acquisition and calcification (Smith and
Roth 1979, Borowitzka 1981, Gao et al. 1993). As for
phytoplankton, there are a wide range of responses
by macroalgae to elevated CO2 concentration. A
52% increase in growth in response to a doubling
of pCO2 was observed for Lomentaria, a species that
uses only CO2 (Kübler et al. 1999). This increase in
growth rate is consistent with the idea that algae
without CCMs are likely to respond to increased
pCO2. Some macroalgal studies have suggested a
negative effect of acidification on particular species,
while other species show positive or no response to
elevated CO2 (Israel et al. 1999, Israel and Hophy
2002, Swanson and Fox 2007). Tropical macroalgal
assemblages have shown positive influences of ele-
vated CO2 on recruitment of noncalcifying macroal-
gae, while inhibiting recruitment of corallines
(Kuffner et al. 2008). It is not clear if these differ-
ences in response were due to reduced survivorship
or competitive ability of calcifying recruits, and ⁄ or
increased competitive ability of noncalcifying algae,
or other factor(s) (Kuffner et al. 2008). Hall-Spen-
cer et al. (2008) demonstrated that within 120 m of
cold CO2 vents (average pH 7.83), macroalgal com-
munities are dominated by fleshy seaweeds, whereas
calcareous seaweeds dominated farther from the
vent (average pH 8.14).

Methods: an appraisal. There are clearly a range
of biological responses to pH manipulation treat-
ments. This observed spectrum of responses may be
due to inherent differences in algal physiology, the
environment in which the algae have grown prior to
experiments (e.g., light climate), the timescale of
the physiological response measured (e.g., short-
term estimates of photosynthesis vs. integrated
growth), duration of the study (days vs. months), or
time of year (Tortell et al. 2008). Another key influ-
ential factor could be the method of pH manipula-
tion because CO2 bubbling and HCl ⁄ NaOH affect
carbonate chemistry differently (see below); there
have been heated discussions on which method is
most suitable (Iglesias-Rodriguez et al. 2008a,b,
Riebesell et al. 2008). Critically, it is difficult to tease
apart the relative importance of each of these
potentially influential factors on the outcome of
experiments. Here, we focus on the different ways
in which carbonate chemistry is altered during pH
manipulation experiments and how this might

1242 C A T R I O N A L . H U R D E T A L .

T
a
b
l
e
2
.

C
a
lc

a
re
o
u

s
a
lg

a
e
.

E
x
a
m
p
le
s
o

f
e
x

p
e
ri
m
e
n
ts
co
n
d
u
ct
e
d
o
n
e
ff
e
ct
s
o
f
C
O
2
co
n
ce
n
tr
a
ti
o
n
o
n
a
lg
a
l
g
ro
w
th

,
ca

lc
ifi

ca
ti

o
n
,
o
r
m
e
ta
b
o

li
c

ra
te

s.
U

n
it
s
fr
o
m
th
e
o
ri
g
in
a
l
p
a
p

e
rs

a
re

re
p

o
rt
e
d
.
S
p
e
ci
e
s
⁄s
ys
te
m
M
e
th
o
d
s
T
im
e
fr
a
m
e
p
H
⁄p
C
O
2
tr
e
a
tm
e
n
ts
R
e
p
li
ca
ti
o
n
R
e
su
lt
s
A
u
th
o
rs
P
h
yt
o
p
la
n
k
to
n
C
o

cc
o

li
th

o
p
h
o

re
s

E
m
il
ia
n
ia
h
u
xl
ey

i
a
n

d
G

ep
h
yr

oc
a
p
sa

oc
ea

n
ic
a
H
C
l⁄
N
a
O
H
,
d
il
u
te
b
a
tc
h
cu
lt
u
re
,
3
0
–
1
5
0
l
m
o
l
p
h
o
to

n
s

Æ
m

)
2

Æ
s)

1
,
1
6
:8
o
r
2
4
:0
L
😀
,
1
5
�C
,
1
0
0
l
M
n
it
ra
te
,
6

.5
l

M
p
h
o
sp
h
a
te
,
f
⁄2

tr
a
ce

e
le
m
e
n
ts
.

A
cc

li
m
a
ti
o
n
fo
r

7

–
9

g
e
n

e
ra
ti
o
n
s
th
e

8
ce

ll
d

iv
is

io
n

s
d

u
ri

n
g
e
x
p
e
ri
m
e
n
t.
p
C
O
2
=
2
8
0
–
7
5
0

.
n

=
3

R
e
d

u
ce

d
ca

lc
ifi
ca
ti
o
n
o
f
1
5

.7
%

in
E

.
h
u

xl
ey
i
a
n

d
4

4
.7

%
in

G
.

oc
ea
n
ic
a
.

M
a
lf

o
rm

e
d
co
cc
o
li

th
s,

sl
ig

h
t

in
cr
e
a
se

in
p

h
o

to
sy

n
th

e
si
s
in
b
o

th
.

R
ie

b
e
se

ll
e
t

a
l.
(2
0
0
0
)
N
o

rt
h

P
a
ci
fi
c
p
h
yt
o
p
la
n
k
to
n
a
ss
e
m
b
la

g
e
s

B
u
b
b
li
n
g
w
it
h
C
O

2
a
ir

m
ix

tu
re
s
o

r
H

C
l⁄

N
a
O

H
,

sh
ip

b
o
a
rd
,
3
0
%

in
ci

d
e
n

t
ir

ra
d

ia
n

ce
,

1
3
�C
,
n
u
tr
ie
n
t
co
n
ce
n
tr
a
ti
o
n
s
n
o
t
re
p
o
rt
e
d
.

1
.5

–
9

d
p

C

O
2
=
2
5
0

,
p

H
=

8
.2

a
n

d
p
C

O
2
=
8
0
0
,
p

H
7

.7
5

n
=

3
S

ig
n

ifi
ca
n
t
re
d
u
ct
io
n

s
(3

6
%

–
8

3
%

)
in

ca
lc

ifi
ca
ti
o
n
in

4
in

d
e
p

e
n
d
e
n

t
e
x

p
e
ri
m
e
n
ts
.
N
o
rt
h

A
tl

a
n

ti
c

p
h
yt
o
p
la
n
k
to
n
a
ss
e
m
b
la
g
e
s
fr
o

m
su

rf
a
ce
a
n

d
d

e
p
th
H
C
l⁄
N
a
O
H
,
sh
ip
b
o
a
rd
,
3
5
0
a
n

d
5

0
l
m
o
l
p
h
o
to
n
s
Æ
m
)
2
Æ
s)
1
,
a
m
b
ie
n
t
te
m
p
e
ra
tu
re
,
n
u
tr
ie
n

t
co

n
ce
n
tr
a
ti
o
n
s
n
o
t
re
p
o
rt
e
d
.

2
h

in
cu

b
a
ti

o
n

3
,

1
0
,
3
6
,
a
n

d
9

1
l

M
C

O
2
n
=
3
S
ig
n
ifi
ca
n
t
in
cr
e
a
se

in
1
4
C

in
co

rp
o

ra
ti

o
n
w
it

h
3

6
a
n

d
9
1
l

M
tr

e
a
tm
e
n
t
co
m
p
a
re

d
w

it
h
1
0
l
M

.
D

e
cr
e
a
se
fo
r

3
l

M
tr
e
a
tm
e
n
t.

H
e
in

a
n

d
S

a
n

d
-J

e
n

se
n

(1
9

9
7

)
E
m
il
ia
n
ia
h
u
xl
ey

i
(c

a
lc
if
yi
n
g

st
ra

in
)
C
O
2
b
u
b
b
li
n
g
,
se
m
ic
o
n
ti
n
u
o
u
s
cu
lt
u
re

,
5

0
a
n
d
4
0
0
l
m
o
l
p
h
o
to
n
s
Æ
m
)
2
Æ
s)
1
,
2
0

�C
a
n

d
2
4
�C
,
f
⁄2
n
u
tr
ie
n
ts
.
A
cc
li
m
a
ti
o
n

in
tr

e
a
tm
e
n
t
co
n
d

it
io

n
s
fo
r
7
g
e
n
e
ra
ti
o

n
s.

p
C
O
2
=
3

5
7

a
n
d
7
5
0
p
p
m
n
=
3
N
o
e
ff
e
ct
o
f
p
C
O
2
o
n
g
ro
w
th
ra
te
o
r
o
rg
a
n
ic
C
p
e
r
ce
ll
a
t

g
iv

e
n
ir
ra
d
ia
n
ce
a
n

d
te

m
p
e
ra
tu
re
tr
e
a
tm
e
n

t.
P

IC
w

a
s
re
d
u
ce

d
in

th
e
h
ig

h
li

g
h

t
⁄7

5
0
p
p
m
tr
e
a
tm
e
n
t.

F
e
n

g
e
t

a
l.
(2
0
0
8
)
C
a
lc

id
is

cu
s

le

p
to

p
or

u
s
a
n
d
C

oc
cl

it
h
u

s

p
el

a
gi

cu
s
H
C
l⁄
N
a
O
H
,
d
il
u
te
b
a
tc
h
cu
lt
u
re
s,
3
5
0
l
m
o
l
p
h
o
to
n
s
Æ
m
)
2
Æ
s)
1
,
1
6

:8
L

😀
,
2
0

�C
fo

r
C

.
le

p
to
p
or

u
s,

1
7

�C
fo
r
C

.
p
el

a
gi

ca
,

1
0
0
l

M
n

it
ra

te
,

6
.5

l
M
p
h

o
sp

h
a
te

,
f
⁄2
tr
a
ce
e
le
m
e
n
ts
.
A
cc
li
m
a
ti
o
n
in
tr
e
a
tm
e
n
t
co
n
d
it
io
n
s
fo
r
1
0
g
e
n
e
ra
ti
o
n
s.

C
.

le
p
to

p
or

u
s:

6
p
C

O
2
tr
e
a
tm
e
n

ts
fr

o
m
9
8
–
9
2
0

l
a
tm

.
C

.
p
el
a
gi

ca
:

3
tr

e
a
tm
e
n

ts
1

4
9
,
3
4
5
,
a
n
d
9
1
5
l
a
tm
.
N
o
t
cl
e
a
r
C
.
le
p
to
p
or
u
s:

m
a
lf

o
rm
e
d
co
cc
o
li

th
s

a
t
lo
w
a
n

d
h

ig
h

p
C
O
2
tr
e
a
tm
e
n
ts
co
m
p
a
re
d
to
co
n
tr
o

l
(3

4
5
p
p

m
).

C
.
p
el
a
gi

cu
s:

n
o
ch
a
n
g
e

i

n
li

th
m

o
rp

h
o

lo
g

y.
B

o
th

sp
e
ci

e
s:

n
o
e
ff
e
ct
o
f
C
O
2
o
n
p
o
p
u
la
ti
o
n
g
ro
w
th
ra

te
b

u
t

in
cr
e
a
se
d
o
rg
a
n

ic
co

n
te

n
t
o
f
ce
ll

s
u

n
d
e
r
h
ig
h
C
O
2
tr
e
a
tm
e
n
ts
.

L
a
n

g
e
r

e
t
a
l.
(2
0

0
6

)
E
m
il
ia
n
ia
h
u
xl
ey

i
C

O
2
b
u
b
b
li
n

g
,

d
il
u
te
b
a
tc
h
cu
lt
u
re
s,
1
5
0
l
m
o
l
p
h
o
to
n
s
Æ
m
)
2
Æ
s)
1
,
1
2
:1
2
L
😀
,
1

9
�C

,
f
⁄2

m
e
d

iu
m

.

A
lg

a
e
a
cc
li
m
a
te

d
a
t

e
x
p
e
ri
m
e
n
ta
l
co
n
d
it
io
n

s
fo

r
9

g
e
n
e
ra
ti
o
n
s.
C
O

2
=

2
8

0
,

3
0
0
,
4
9
0
,
6
0
0
,
7
5
0
p
p
m
v
n
=

6
D

o
u
b
li
n
g
in
p
a
rt
ic
u
la
te

in
o

rg
a
n
ic
ca
rb
o

n
a
n

d
p
a
rt
ic
u
la
te
o
rg
a
n
ic
ca
rb
o

n
,

a
n
d
d
e
cr
e
a
se
in
g
ro
w
th
ra
te
s
a
t
7
5

0
co

m
p
a
re
d
to
2
8
0
p
p
m

v.
N

o
e
ff

e
ct
o
f
d
o
u
b

le
d

C
O
2
o
n
li
th
m
o
rp
h
o
lo
g
y.

Ig
le

si
a
s-

R
o

d
ri

g
u

e
z

e
t
a
l.
(2
0
0
8

a
)

O C E A N A C I D I F I C A T I O N A N D A L G A L M E T A B O L I S M 1243

T
a
b
l
e
2
.
C
o
n
ti
n
u
e
d
.
S
p
e
ci
e
s
⁄s
ys
te
m
M
e
th
o
d
s
T
im
e
fr
a
m
e
p
H
⁄p
C
O
2
tr
e
a
tm
e
n
ts
R
e
p
li
ca
ti
o
n
R
e
su
lt
s
A
u
th
o
rs
M
a
c
ro
a
lg
a
e

C
or

a
ll

in
a

p
il

u
li

fe
ra

C
o
n
st
a
n
t
a
e
ra
ti
o
n
w
it
h
C
O
2
⁄a
ir
m
ix
,
1
5

cm
h

ig
h
w
it

h
1

4
cm

d
ia

m
e
te

r
2
0
�C
,
3
0
l
m
o
l
p
h
o
to
n
s
Æ
m
)
2
Æ

s)
1
.

1
2

h
p
C

O
2
=
3
5
0
,
p

H
8

.2
a
n

d
p
C
O
2

1
,5

0
0
a
n
d
p
H
7
.6
P
se
u
d

o
–

re
p
li
ca
te
d

4
4

%
re

d
u
ct
io
n
in
ca
lc
ifi
ca
ti
o
n
.

G
a
o

e
t
a
l.
(1
9

9
3

)

C
ru

st
o

se
C

or
a
ll

in
e

re
cr

u
it

s
In
je
ct
io
n
o

f
1

0
%
H
C

l⁄
fr

e
sh

w
a
te

r
m

ix
a
t

in
fl

o
w

to
5

0
0

L
o

u
td

o
o

r
m

e
so

co
sm

s,
a
m

b
ie
n
t
ir
ra
d
ia
n
ce
,
se
a
w
a
te
r
te
m
p
e
ra
tu
re
,
a
n
d
n
u
tr
ie
n
ts
.

7
w

e
e
k
s
C
o
n
tr
o

l
p
C

O
2

=
4

0
0
l
a
tm
,
p
H
8

.1
7

a
n

d
tr

e
a
tm
e
n

t
p
C

O
2

=
7

6
5

l
a
tm
,
p
H
7

.9
1

a
t
m
id
d
a
y.

D
iu

rn
a
l

fl
u

ct
u

a
ti
o
n
s
o

f
p

H
o

f
�

0
.5
u
n

it
s.

n
=

3
M

a
rk
e
d
re
d
u
ct
io
n
in
re

cr
u

it
m

e
n

t
a
n

d
%

co
ve

r
o

f
cr

u
st

o
se

c

o
ra

ll
in

e
s

in
h

ig
h

–
C

O
2
tr
e
a
tm
e
n

t.
In

cr
e
a
se

in
n

o
n
ca
lc
if
yi
n
g

m
a
cr

o
a
lg
a
e
.

K
u

ff
n

e
r
e
t
a
l.
(2
0
0
8
)
P
or

ol
it

h
on

on
ko

d
es

C
O
2
b
u
b
b
le
d
in
to
m

ix
in

g
ta

n
k

s,
fe

d
in

to
2

0
L

a
q

u
a
ri

a
,
a
m
b
ie
n
t
ir
ra
d
ia
n
ce
7
0

0
–
1

,2
0
0
l
m
o
l
p
h
o
to
n
s
Æ
m
)
2
Æ
s)
1
,

tw
o

te
m
p
e
ra
tu
re
tr
e
a
tm
e
n

ts
:

2
5

�C

–
2

6
�C

a
n
d
2
8
�C
–
2
9
�C
,
a
m
b
ie
n
t
n
u
tr
ie
n
ts
.

8
w

e
e
k
s
C
o
n
tr
o
l
p
H
=
8
.0
–
8

.4
(3

8
0
p
p

m
v)

;
tr

e
a
tm
e
n

ts
=

p
H

7
.8

5
–
7

.9
5

(C
O

2
5
2
0
–
7
0
5
p
p
m
v)
a
n
d
p
H
7

.6
0

–
7

.7
0

(1
,0

0
0
–
1

,3
0

0
p
p
m

v)
.

n
=

5
B

le
a
ch

in
g

,
m

a
rk
e
d

d
e
cr

e
a
se
in
n
e
t
p
h
o
to
sy
n

th
e
si

s
a
n

d
ca
lc
ifi
ca
ti
o
n
w
it

h
d

e
cr
e
a
si
n
g
p
H
tr
e
a
tm
e
n
t.
A
n
th
o

n
y

e
t
a
l.
(2
0
0
8
)

S
u

b
ti

d
a
l

a
n
d
in

te
rt

id
a
l

co
m

m
u
n
it

ie
s

‘‘
N

a
tu
ra
l

la
b

o
ra

to
ry

’’
(i

.e
.,

a
co

ld
-w

a
te

r,
su

b
m

a
ri
n
e
C
O
2

ve
n

t)
.

C
o

m
m

u
n

it
y

su
rv

e
ys

o
f
sp
e
ci
e
s

p
re

se
n

ce
a
n

d
%
co
ve

r
a
t

va
ri

o
u
s
d

is
ta

n
ce
s
fr
o
m
C
O
2
ve
n
ts
.
p
H

g
ra

d
ie

n
t
o
f

8
.1

4
–
6

.5
7

R
e
p

li
ca
te
d

%
co

ve
r

o
f
ca
lc
a
re
o
u

s
se

a
w
e
e
d
s
d
e
cr
e
a
se

d
cl

o
se

r
to

ve
n
ts
,
%
co
ve
r
n
o
n
ca

lc
a
re

o
u
s
se
a
w
e
e
d
s
in
cr
e
a
se
d
.

H
a
ll

-S
p

e
n

ce
r

e
t
a
l.
(2
0
0
8
)

1244 C A T R I O N A L . H U R D E T A L .

impact the physiological response of algae. We then
consider the steps required to design and imple-
ment improved experiments on ocean acidification,
and finally, we discuss the implications of a lack of
standardization in the other environmental factors
on making a direct comparison between the results
from different published studies.

Comparison of HCl additions versus CO2 bubbling on
seawater carbonate chemistry: The most commonly
used means to simulate the effects of acidification
involves an initial adjustment of the CO2 speciation
of a seawater culture medium to achieve the desired
degree of acidification and then maintaining those
conditions as growth of the alga(e) proceeds. There
are two methods of decreasing seawater pH, by bub-
bling the seawater with CO2 gas or by the addition
of acid (commonly HCl).

Each method produces a specific target CO2 spe-
ciation, but they achieve this in different ways
(Langdon 2003). The equilibration of CO2 gas with
seawater fixes the equilibrium pCO2 to that of the
control gas, without any change in total alkalinity,
AT. As a result, CT will increase from its initial value.
In contrast, addition of HCl to seawater involves a
known decrease in the original AT, but no change
in CT.

Figure 1 shows simulated values of pCO2,
[HCO3

)], [CO3
2)], and calcite saturation (X)

achieved by acidifying a seawater sample of typical
surface-water composition using each method.
There is very little difference in the two methods
for a given pH with respect to [CO3

2)] and X, but
at pH 7.5, pCO2 and HCO3

) are 23% and 22%
lower, respectively, using HCl additions compared
to CO2 bubbling. This is because when seawater is
acidified with HCl, CT remains constant (provided

CO2 does not escape into the gas phase), and so
[HCO3

)] increases slightly as a small amount of
CO3

2) is converted to HCO3
) by reaction with H+

ions. However, when CO2(g) is added, CT is
increased, mainly in the form of HCO3

).
CO2 bubbling is arguably much closer to the

actual ocean acidification that is currently affecting
the surface ocean. Here, we suggest a third method
that will achieve the same effect without bubbling
with CO2, which is by adding, separately, equivalent
concentrations of HCl and sodium bicarbonate solu-
tions. In this case, the NaHCO3 exactly neutralizes
the alkalinity decrease caused by the HCl and sup-
plies the increased CT (mostly as HCO3

)) that
would be the result of CO2 gas bubbling. While this
method has not been used in published biological
experiments to date, the use of separate HCl and
NaHCO3 additions has the advantage that it avoids
any possible physiological effects that bubbling itself
might cause (see later). The changes in CO2 specia-
tion induced by this method are identical to those
achieved by bubbling with CO2 gas, as shown in
Figure 1.

How might the method of pH manipulation affect algal
physiology? The question is whether the different
methods that have been used to manipulate seawa-
ter pH could affect the outcome of experiments.
For algae without CCMs that rely on diffusive
uptake of CO2, the 23% difference in H2CO3

*

between the CO2-bubbling and the HCl-addition
methods at pH 7.5 (and indeed a 17% difference at
pH 7.75; Fig. 1) could cause significantly lower rates
of photosynthesis and growth with the HCl method.
For noncalcifying algae with CCMs, the concentra-
tion of CO2 at the alga’s surface controls the activity
of the CCM and carbonic anhydrase (Giordano
et al. 2005); the 23% higher H2CO3

* concentration
(with CO2 bubbling) may be sufficient to cause a
down-regulation of the CCM and CA relative to the
HCl method. Such a down-regulation in the CCM
and the CA could result in the reallocation of cellu-
lar energy to, for example, more rapid growth.

At the current oceanic pH of �8.1, HCO3)
concentrations in seawater are sufficient to saturate
photosynthesis in most organisms with CCMs
(Giordano et al. 2005). Therefore, for algae with
CCMs that utilize bicarbonate, the 22% difference
that results from the two methods of pH manipula-
tion should not affect photosynthesis per se because
HCO3

) is not limiting. The saturation state of cal-
cite (X) is key in determining if organisms can lay
down calcium carbonate, and this is similar for both
methods of pH manipulation (Fig. 1). Whether the
22% difference in HCO3

) concentration between
the two methods might influence calcification is not
clear.

A direct comparison of the two methods (CO2
bubbling or HCl ⁄ NaOH additions) of CO2 manipu-
lation on the responses of algae has not been made.
However, the responses of calcifying strains of coc-

Fig. 1. Calculated equilibrium CO2 parameters as a function
of seawater pH for a seawater sample having initial
CT = 1,900 lmol Æ kg

)1 and AT = 2,300 lmol Æ kg
)1, salinity 35

and temperature 10�C, that has been acidified either with strong
acid HCl or CO2 gas: (a) equilibrium CO2 partial pressure pCO2,
(b) bicarbonate ion concentration [HCO3

)], (c) carbonate ion
concentration [CO3

2)], and (d) saturation ratio X for calcite. It is
assumed that the HCl addition involves insignificant dilution of
the seawater.

O C E A N A C I D I F I C A T I O N A N D A L G A L M E T A B O L I S M 1245

colithophores to ocean acidification have been
tested (by different research groups and often,
when the same species was employed, using differ-
ent strains) using the two methods. Feng et al.
(2008) used the CO2-bubbling method to show that
the particulate inorganic carbon (PIC) production
rate of E. huxleyi remains unaffected by high CO2
concentration at low irradiances for growth, but it is
decreased by high CO2 concentration at high
growth irradiances. The particulate organic carbon
(POC) per cell increased in high CO2(aq) at the
higher growth irradiances. Iglesias-Rodriguez et al.
(2008a) also used the CO2-bubbling method on
E. huxleyi, working at light saturation, and deter-
mined that while increased CO2 concentrations
increased the POC and PIC per cell, there was no
change in the PIC:POC ratio with increasing CO2,
and a significant decrease in the specific growth rate
of the cells at the highest CO2 concentration used.

Using the HCl ⁄ NaOH method of CO2 manipula-
tion, Riebesell et al. (2000) showed that high CO2
concentration decreased the rate of calcification
(on a per cell basis) in E. huxleyi and Gephyrocapsa
oceanica, with increased calcification at lower than
present CO2 levels. Again, using the HCl ⁄ NaOH
technique, Langer et al. (2006) observed no
increase in inorganic carbon precipitation by Co.
pelagicus by increasing CO2 concentrations from that
occurring during the last glacial maximum level to
more than twice present-day CO2 levels, while Ca.
leptoporus had the highest rate of calcification at the
present CO2 level, with lower rates at more than
twice the present-day level and an even greater
decrease at the last glacial maximum level. Clearly,
there are intergeneric variations in the response of
calcification to CO2 concentrations in all experi-
ments using the HCl ⁄ NaOH method. Whether the
difference in response of E. huxleyi to CO2 variations
reported by Riebesell et al. (2000), Feng et al.
(2008), and Iglesias-Rodriguez et al. (2008a) is a
function of the different methods of changing CO2
or of other experimental differences or the use of dif-
ferent algal strains awaits further experimentation.

Methodological considerations: Here we discuss
eight points that we consider essential in designing
experiments that examine the impacts of ocean
acidification on algae. Most considerations are rele-
vant to studies involving either macroalgae or micro-
algae.

1. Which method of manipulating pH to choose? This
will depend on the research question being asked.
If the question is related to carbon acquisition and
rates of photosynthesis and growth, then adding
CO2 to shift seawater pH most closely mimics the
changes in seawater carbon speciation predicted in
the future due to climate change (changing CT,
constant alkalinity). For phytoplankton, however,
there may be side effects of directly bubbling with
CO2, such as damage to fragile phytoplankton
through the effects of small-scale turbulence with

increasing sensitivity in the order green algae >
cyanobacteria > diatoms > dinoflagellates (Thomas
and Gibson 1990). The HCl method is relatively
simple to implement and can be used for investigat-
ing the effects of changes in pH, H2CO3

*, or CO3
2)

as long as the resulting changes in AT (and lack of
change in CT) are of no concern. In addition, the
scale of experiment might affect the pH manipula-
tion used. For small-scale laboratory cultures and
shipboard experiments, pH can be manipulated
using CO2 bubbling, but the HCl-addition tech-
nique may be more practicable for large-scale meso-
cosm experiments (Kuffner et al. 2008). Whichever
method is chosen, clear interpretation of physiologi-
cal responses of algae to pH manipulation will be
facilitated by a thorough understanding of the
entire carbonate system in the culture medium. To
this end, it is essential to conduct regular monitor-
ing of at least two of the carbonate analytical param-
eters (see below).

2. Methods of pH control. For CO2 bubbling, pertur-
bations can take place using pure CO2 gas (Leclercq
et al. 2000, Israel and Hophy 2002), CO2 ⁄ air mix-
tures (Gao et al. 1993, Riebesell et al. 2000, Engel
et al. 2005), or CO2 mixed with other gases (O2,
N2) (the latter to permit an analysis of CO2 ⁄ O2
interaction effects on growth, Kübler et al. 1999).
CO2 ⁄ air and gas mixtures at certified concentrations
are expensive, and using gas mixers to produce
appropriate CO2 ⁄ gas mixtures is an option (Kübler
et al. 1999, Engel et al. 2005). Parsons et al. (1992)
describe how to build an inexpensive gas-mixing sys-
tem that was used effectively to grow a macroalga at
a range of pCO2 (Kübler et al. 1999). Careful bub-
bling of small amounts of inexpensive food-grade
100% CO2 while monitoring pH is also effective in
providing required pH ⁄ CO2 concentration (C. D.
Hepburn, K. Currie, and C. L. Hurd, unpublished
data).

When using HCl, it is important not to introduce
any other perturbation to the system. The HCl
should be made up in NaCl solution such that the
total ionic strength is similar to that of the seawater,
that is, 0.7 M (e.g., 0.1 M HCl and 0.6 M NaCl).
The HCl solution should not be contaminated with
trace metals if this is important (e.g., in iron-limited
waters).

Higher pH and reduced pCO2 (i.e., mimicking
the preindustrial era) can also be achieved through
bubbling seawater with another gas (e.g., N2), CO2-
depleted air (produced by pumping ambient air
through Na2CO3 traps) (Leclercq et al. 2000, Engel
et al. 2005), or CO2-depleted gas mixtures (Kübler
et al. 1999). Due to seawater’s strong buffering
capacity, bubbling with CO2-free or CO2-depleted
gases can take some time and a significant amount
of depleted gas before the required pCO2 is
reached. Shifting pH using concentrated NaOH is a
faster and straightforward method, but, as for HCl,
the resulting carbonate speciation is different from

1246 C A T R I O N A L . H U R D E T A L .

that which occurs in situ in response to changing
pCO2.

3. Monitoring pH and carbonate analytical parameters
during experiments. Algal metabolism will alter the
pH (and thus carbon speciation) of the incubation
medium. It is essential to measure at least daily pH
and one other carbonate analytical parameter of the
incubation medium. Measurement of two of the car-
bonate analytical parameters (pH, CT, alkalinity, or
pCO2) allows determination of the other two param-
eters and the concentrations of the species when
combined with knowledge of the carbonate equilib-
rium constants (Lewis and Wallace 1998, Hunter
2007). pH measurements can be made either poten-
tiometrically, using high precision and carefully
maintained electrode ⁄ meter combinations, or opti-
cally, using spectrophotometric measurement of a
dye ⁄ seawater mixture (Tapp et al. 2000, Ohline
et al. 2007). Care must be taken not to lose (or
gain) CO2 by exchange with the atmosphere during
the subsampling and measurement process. Careful
calibration of the pH measurement system is
required using appropriate buffer solutions made
up in synthetic seawater (Dickson 1993a,b, Dickson
et al. 2007). The defined pH value of the buffer is
temperature-dependent, so careful temperature con-
trol is required. The measurement of other carbon-
ate analytical parameters is described by Dickson
et al. (2007), and pCO2 can be measured directly
using membrane inlet mass spectrometry (MIMS;
Gueguen and Tortell 2008).

4. Modification of seawater pH ⁄ pCO2 within an experi-
mental system. Modification of pH ⁄ pCO2 should
occur in separate (mixing) containers, not directly
in seawater in contact with the study organisms.
This can be achieved in flow-through systems by
adding the amendment solution or gas to indepen-
dent header tanks (e.g., collapsible bags to prevent
gas exchange with headspace), to tubing (Kuffner
et al. 2007), or in mixing chambers (Leclercq et al.
2000) upstream of culture containers in flow-
through systems. The bubbling of CO2 ⁄ air mixtures
directly into culture containers is also acceptable for
macroalgae (as long as it is at the correct pCO2 for
the treatment) but could damage delicate phyto-
plankton (Thomas and Gibson 1990, Berdalet et al.
2007). Direct addition of acid or concentrated CO2
into the culture medium surrounding experimental
subjects makes it difficult to separate damage due to
direct shock to the organisms from localized spikes
of pH from the cumulative effect of altered
pH ⁄ pCO2 due to the treatment.

5. Biological side effects and pH range during experi-
ments. Photosynthesis (increasing pH), respiration,
and calcification (lowering pH) by algae can
strongly modify the pH and carbonate chemistry of
seawater in culture containers (Israel and Hophy
2002). Such effects are particularly important when
using macroalgae that are often large and have
rapid metabolic rates, and in mesocosms where bio-

mass levels, and hence biological activity, can be
high. For experiments that require a constant pH,
variation due to photosynthesis or respiration can
be reduced by high seawater to macroalgal tissue
ratios and ⁄ or systems that have seawater flow from
reservoirs with fixed pH levels, and for microalgae,
continuous and semicontinuous culture systems can
be advantageous over batch cultures because of the
constant replenishment with fresh media. Critically,
care must be taken not to mistake the effects of
short periods of unrealistic pH resulting from meta-
bolic processes to the effect of different pCO2 and
pH treatments that simulate acidification.

Seawater pH naturally varies on timescales from
diurnal to seasonal. An understanding of this natu-
ral variability is important when designing an
experiment and interpreting results. Culture cham-
bers with inflow from surrounding coastal waters
can exhibit pH fluctuations of up to 0.6 or 1 pH
unit due to natural diurnal variations in the seawa-
ter source (Swanson and Fox 2007, Anthony et al.
2008). Prior to experiments that use seawater
pumped from coastal waters, seawater pH should
be monitored at least over a daily cycle so that nat-
ural fluctuations experienced by the algae are
known. In experiments to examine community-level
effects of pH, the goal may be to achieve a variable
pH that reflects that of the natural environment.
For example, a diurnal cycle of �0.6 pH units was
evident for mesocosm cultures of coralline algae
(Kuffner et al. 2007), while seawater pH increased
by 0.2–0.3 units gradually over the first 11 d of the
21 d long PeECE III mesocosm experiments (Bell-
erby et al. 2008).

6. Overcoming chemical artifacts. In both methods
for simulating acidification, secondary changes in
CO2 speciation are possible through the following
processes. For calcifying algae, the most important
will be dissolution of biogenic CaCO3 in the experi-
mental chamber as a result of acidification. This
phenomenon will increase both CT and AT. How-
ever, in a realistic experimental setup, one would
want to know about dissolution of biogenic CaCO3
as an outcome, so it is likely that this would be mon-
itored, either by measuring weight loss of the
CaCO3 or by parallel measurements of any two of
the CO2 system parameters. For example, one could
monitor pH or pCO2 continuously during the cul-
ture experiment and also take samples for CT
and ⁄ or AT measurements. This approach would
enable any changes arising from CaCO3 dissolution
to be corrected for.

Another secondary effect is loss of CO2 gas
because the equilibrium of pCO2 in the chamber is
greater than that of the ambient atmosphere. For
culture experiments controlled by addition of CO2
gas, this can be minimized by partly enclosing the
ambient air so that both air and water phases
remain in equilibrium. The disadvantage of this
approach is that each chamber requiring a different

O C E A N A C I D I F I C A T I O N A N D A L G A L M E T A B O L I S M 1247

pCO2 condition must be supplied with its own stan-
dard air-CO2 mixture.

For a system maintained by periodic additions of
HCl, the extent of CO2 loss can be monitored by
measurement of CT before and after. In practice,
the loss of CO2 under realistic pH conditions
(pH > 7.5) does not appear to be very large if the
culture chamber is fitted with an inflated plastic bag
that has a volume similar to, or smaller than, the
volume of water in the chamber. This is because the
quantity of CO2 in a head space is extremely small
compared to that contained in an equivalent vol-
ume of seawater.

7. Replication. A problem with some studies on
ocean acidification is that they exhibit low levels of
replication and ⁄ or that replicates are not truly inde-
pendent of each other; this is especially the case for
macroalgae (Tables 1 and 2). Care must be taken to
provide independent replicates required for the cor-
rect application of statistical tests (i.e., each repli-
cate culture tank should have its seawater modified
to the appropriate pH independently, not in one
header tank per treatment). Psuedoreplication (i.e.,
growing ‘‘replicate’’ algae in the same treatment
container) must be avoided (see Hurlbert 1984).
Obtaining appropriate levels of independent repli-
cation is especially difficult for the larger macroal-
gae (e.g., Fucales and Laminariales), which can be
problematic to maintain in culture long-term, and
enclosing individuals or populations for field manip-
ulations is extremely difficult. Experiments using
macroalgae naturally suffer from high levels of stan-
dard deviation between replicates because each rep-
licate comprises one individual (compared to a
phytoplankton culture of millions of cells); a popu-
lation response of macroalgae to a treatment is
therefore difficult to achieve.

8. Other environmental factors. A preoccupation
with mimicry of seawater carbonate chemistry may
result in overlooking other factors important for
algal growth (e.g., temperature, UV radiation, light
climate [photon flux density, light:dark cycle], nutri-
ent concentrations, water motion). It is evident
from Tables 1 and 2 that these factors vary widely
across the range of studies considered. Rates of
calcification and carbon acquisition are energy (i.e.,
light) dependent. Light limitation may result in
algae taking up CO2 in preference to HCO3

); light
levels that are too high might induce photoinhibi-
tion and cause redirection of energy to cellular
repair mechanisms and away from growth. Suitable
irradiances can be determined from the results of
photosynthesis versus irradiance curves. Tempera-
ture influences all aspects of algal growth and physi-
ological rates. Inadequate water motion will cause
diffusion-limited growth, particularly for macroalgae
(Hurd 2000). Ideally, standardization between
studies would be valuable and would permit direct
comparison of the results from different studies.
There is also a need for experiments that test the

interactive effects of ocean acidification and other
predicted changes in climate (Feng et al. 2008). We
recommend providing experimental conditions that
are as similar to those in the natural environment
as possible. Nevertheless, it is essential to report all
the experimental growth conditions listed above to
permit critical evaluation of the results.

Incubation timescales and prior conditioning of organ-
isms: Several important issues are apparent from
the CO2-manipulation experiments conducted to
date. A major issue is the degree of physiological
response to altered environmental conditions, and
this is strongly influenced by the timescale of exper-
iments. The outcomes of these studies have mainly
been at the level of acclimation, as defined by Raven
and Geider (2003), that is, studies of organisms that
have had time to show qualitative or quantitative
changes in gene expression during growth in
response to the experimental treatments. An impli-
cit or explicit assumption in interpreting the results
is that the experiments did not last long enough to
permit adaptation (genetic change) of the a strain
in unialgal cultures of the kind investigated over
1,000 generations of the freshwater alga Chlamydo-
monas reinhardtii by Collins and Bell (2004) using
the CO2-enrichment method. ‘‘Natural laborato-
ries’’ such as underwater CO2 vents also provide
opportunities to study adaptation (Hall-Spencer
et al. 2008). Very short-term experiments (e.g., mea-
surements of short-term inorganic 14C assimilation,
net oxygen exchange, or chlorophyll-fluorescence-
derived electron transport rates), involving exposure
of organisms grown at the present day, or some
other CO2 concentration to step changes in CO2,
only permit the operation of cellular mechanisms
termed regulation (preexisting metabolic machin-
ery; Raven and Geider 2003) but are essential in
defining the kinetic properties of the inorganic car-
bon acquisition mechanisms under a given set of
growth conditions. Short-term (2 h) photosynthesis
experiments on natural marine phytoplankton
assemblages do not give time for complete acclima-
tion to new experimental conditions and may (as
acknowledged by Hein and Sand-Jensen 1997, see
also Schippers et al. 2004) overestimate the longer-
term (days) effect of the increased CO2 on meta-
bolic rates.

Research using laboratory cultures suffers from
selection of genotypes favored by the maintenance
conditions for the isolate. These conditions include
the absence of UV radiation, low PAR fluxes, and
unnatural nutrient solute concentrations (including
inorganic carbon) in the medium if the isolates have
been in culture for a long time (months to years).
Against this, there is the possibility of using data from
other experiments involving the same algal strain in
planning and interpreting experiments for the organ-
isms that were isolated and cultured a long time ago.
In some cases, laboratory experiments used recently
isolated strains when these were available: Burkhardt

1248 C A T R I O N A L . H U R D E T A L .

et al. (1999) used Asterionella glacialis, Coscinodiscus
wailesii, Thalassiosira punctigera, and Scrippsiella
trochoidea strains that had recently been isolated from
the North Sea but obtained Phaeodactylum tricornutom
from a culture collection. For mesocosms, there is
the advantage that the algae examined have not spent
a long period in culture, but there is the problem of
separating the main species contributing to the algal
biomass from other species if more than cell counts
are needed, and the requirement to check by molecu-
lar phylogenetic means that the same genotype is
involved not only in the various CO2 treatments but
also within replicates of a given treatment. This prob-
lem of intraspecific genotypic variability also applies
to the use of macroalgae and seagrasses from natural
populations for CO2 manipulations in the field or in
the laboratory. For natural vents, the presence of
chemicals in addition to CO2 in the vent fluids, plus
the problem of advection of parcels of water and
their contained biota to and away from the vent site,
are complicating factors for planktonic organisms
(Dando et al. 2000), and of genotypic selection of
adjacent macroalgae and seagrasses.

In conclusion, there is a need to run experi-
ments using both approaches to altering CO2
chemistry, so that any differences in approaches as
a contributing factor to the wide range of
responses reported for both micro- and macroalgae
upon alteration of CO2 chemistry can be
accounted for. It is essential that the carbon specia-
tion within culture vessels is carefully monitored at
least daily during incubation experiments by mea-
suring pH and one other analytical parameter (CT,
alkalinity or pCO2) of the seawater carbonate sys-
tem. There is also a need to better standardize
across the scientific community the timescales of
preconditioning of samples (e.g., natural communi-
ties vs. laboratory cultures) and of incubations dur-
ing experiments (from 2 h, Schippers et al. 2004,
to 2 years, Collins and Bell 2004, 2006). Finally,
independent replication of experimental pCO2
treatments is an essential prerequisite for statisti-
cally meaningful results, and other incubation con-
ditions should mimic natural environmental
conditions (e.g., light climate, inorganic nutrient
concentrations) wherever possible.

This work was funded by University of Otago Research Grants
to C. L. H., C. D. H., and K. A. H., and a Royal Society of
New Zealand ISAT-linkages grant to C. L. H. J. A. R.’s work
on calcified algae is supported by the Natural Environment
Council (UK). The University of Dundee is a registered Scot-
tish charity, No. SC015096. We thank Philip Boyd for his
insightful comments, and Daniel Pritchard and Christopher
Cornwall for helpful discussions. We are grateful to three
anonymous reviewers for their perceptive and generous
reviews. This manuscript is dedicated to our colleague and
mentor Prof. Peter Bannister.

Adey, W. H. 1998. Coral reefs: algal structured and mediated eco-
systems in shallow, turbulent, alkaline waters. J. Phycol. 34:393–
406.

Anthony, K. R. N., Kline, D. I., Diaz-Pulido, G., Dove, S. & Hoegh-
Gludberg, O. 2008. Ocean acidification causes bleaching and
productivity loss in coral reef builders. Proc. Natl. Acad. Sci.
U. S. A. 105:17442–6.

Balch, W., Drapeau, D., Bowler, B. & Booth, E. B. 2007. Prediction
of pelagic calcification rates using satellite measurements.
Deep-Sea Res. Part II Top. Stud. Oceanogr. 54:478–95.

Barcelos e Ramos, J., Biswas, H., Schulz, K. G., La Roche, J. &
Riebesell, U. 2007. Effect of rising carbon dioxide on the
marine nitrogen fixer Trichodesmium. Glob. Biogeochem. Cycles
21:GB2028.

Bellerby, R. G. J., Schulz, K. G., Riebesell, U., Neill, C., Nondal, G.,
Heegaard, E., Johannssen, T. & Brown, K. R. 2008. Marine
ecosystem community carbon and nutrient uptake stoichiom-
etry under varying ocean acidification during the PeACE III
experiment. Biogeosciences 5:1517–27.

Berdalet, E., Peters, F., Koumandou, V. L., Roldan, C., Guadayol,
O. & Estrada, M. 2007. Species-specific responses of dinofla-
gellates to quantified small-scale turbulence. J. Phycol.
439:965–77.

Bijma, J., Spero, H. J. & Lea, D. W. 1999. Reassessing foraminiferal
stable isotope geochemistry: impact of the oceanic carbonate
system (experimental results). In Fisher, G. & Wefer, G. [Eds.]
Use of Proxies in Paleoceanography: Example from the South Atlantic.
Springer-Verlag, Berlin, pp. 498–512.

Borowitzka, M. A. 1981. Photosynthesis and calcification in the
articulated coralline red algae Amphiroa anceps and A. foliacea.
Mar. Biol. 62:17–23.

Borowitzka, M. A. 1987. Calcification in algae: mechanisms and the
role of metabolism. CRC Crit. Rev. Plant Sci. 6:1–45.

Burkhardt, S., Riebesell, U. & Zondervan, I. 1999. Effects of growth
rate, CO2 concentration, and cell size on the stable carbon
isotope fractionation in marine phytoplankton. Geochim. Cos-
mochim. Acta 63:3729–41.

Caldiera, K. & Wickett, M. E. 2003. Anthropogenic carbon and
ocean pH. Nature 425:365.

Chave, K. E., Deffeyes, K. S., Weyl, P. K., Garrels, R. M. & Thomp-
son, M. E. 1962. Observations on the solubility of skeletal
carbonates in aqueous solutions. Science 137:33–4.

Chen, X., Qiu, C. E. & Shao, J. Z. 2006. Evidence for K+-dependent
HCO3

) utilization in the marine diatom Phaeodactylum tricor-
nutum. Plant Physiol. 141:731–6.

Chisholm, J. R. M. 2003. Primary productivity of reef-building
crustose coralline algae Limnol. Oceanogr. 48:1376–87.

Clark, D. R. & Flynn, K. J. 2000. The relationship between dissolved
inorganic carbon concentration and growth rate in marine
phytoplankton. Phil. Trans. R. Soc. Lond. B 267:953–9.

Collins, S. & Bell, G. 2004. Phenotypic consequences of 1000 gen-
erations of selection at elevated CO2 in a green alga. Nature
431:137–45.

Collins, S. & Bell, G. 2006. Evolution of natural algal populations at
elevated CO2. Ecol. Lett. 9:129–35.

Dando, P. R., Aliani, S., Arab, H., Bianchi, C. N., Brehmer, M.,
Cociti, S., Fowler, S. W., et al. 2000. Hydrothermal studies in
the Aegean Sea. Phys. Chem. Earth B 25:1–8.

Dason, J. S., Huertas, I. E. & Colman, B. 2004. Source of inorganic
carbon for photosynthesis in two marine dinoflagellates.
J. Phycol. 40:285–92.

De’ath, G., Lough, J. M. & Fabricius, K. E. 2009. Declining
coral calcification on the Great Barrier Reef. Science
323:116–9.

Dickson, A. G. 1993a. The measurement of pH in seawater. Mar.
Chem. 44:131–42.

Dickson, A. G. 1993b. pH buffers for sea water media based on the
total hydrogen ion concentration scale. Deep-Sea Res. 40:107–
18.

Dickson, A. G., Sabine, C. L. & Christian, J. R. [Eds.] 2007. Guide to
Best Practices for Ocean CO2 Measurements. PICES Special Publi-
cation 3. North Pacific Marine Science Organisation (PICES)
Sidney, British Columbia, Canada, 191 pp.

Doney, S. C., Fabry, V. J., Feely, R. A. & Kleypas, J. A. 2009. Ocean
acidification: the other CO2 problem. Annu. Rev. Mar. Sci.
1:169–92.

O C E A N A C I D I F I C A T I O N A N D A L G A L M E T A B O L I S M 1249

Duggins, D. O., Simenstad, C. A. & Estes, J. A. 1989. Magnification
of secondary production by kelp detritus in coastal marine
ecosystems. Science 245:170–3.

Engel, A., Zondervan, I., Aerts, K., Beaufort, L., Benthien, A., Chou,
L., Delille, B., et al. 2005. Testing the effect of CO2 concen-
trations on a bloom of the coccolithophorid Emiliania huxleyi
in mesocosm experiments. Limnol. Oceanogr. 50:493–507.

Feely, R. A., Sabine, C. L., Lee, K., Berelson, W., Kleypas, J., Fabry,
V. J. & Millero, F. J. 2004. Impact of anthropogenic CO2 on the
CaCO3 system of the oceans. Science 305:362–6.

Feng, Y., Warner, M. E., Zhang, Y., Sun, J., Fu, F.-X., Rose, J. M. &
Hutchins, D. A. 2008. Interactive effects of increased pCO2,
temperature and irradiance on the marine coccolithophore
Emiliania huxleyi (Prymnesiophyceae). Eur. J. Phycol. 43:87–98.

Field, C. B., Behrenfeld, M. J., Randerson, J. T. & Falkowski, P.
1998. Primary production of the biosphere: integrating ter-
restrial and oceanic components. Science 281:237–40.

Frankignoulle, M. & Canon, C. 1994. Marine calcification as a
source of carbon dioxide: positive feedback of increasing
atmospheric CO2. Limnol. Oceanogr. 39:458–62.

Fu, F.-X., Mulholland, M. R., Garcia, N. S., Beck, A., Bernhardt, P.
W., Warner, M. E., Sañudo-Wilhelmy, S. A. & Hutchins, D. A.
2008. Interactions between changing pCO2, N2 fixation, and
Fe limitation in the marine unicellular cyanobacterium Cro-
cosphaera. Limnol. Oceanogr. 53:2472–82.

Fu, F.-X., Warner, M. E., Zhang, Y., Feng, Y. & Hutchins, D. A. 2007.
Effects of increased temperature and CO2 on photosynthesis,
growth, and elemental rations in marine Synechococcus and
Prochlorococcus (cyanobacteria). J. Phycol. 43:485–96.

Gao, K. Y., Aruga, Y., Asada, K. & Kiyohara, M. 1993. Calcification in
the articulated coralline algae Corallina pilulifera, with special
reference to the effect of elevated CO2 concentration. Mar.
Biol. 117:129–32.

Gattuso, J.-P., Allemand, D. & Frankignoulle, M. 1998a. Effect of
calcium carbonate saturation of seawater on coral calcifica-
tion. Glob. Planetary Change 18:37–46.

Gattuso, J.-P., Frankignoulle, M. & Wollast, R. 1998b. Carbon and
carbonate metabolism in coastal aquatic systems. Annu. Rev.
Ecol. Syst. 29:405–34.

Giordano, M., Beardall, J. & Raven, J. A. 2005. CO2 concentrating
mechanisms in algae: mechanisms, environmental modulation
and evolution. Annu. Rev. Plant Biol. 56:99–131.

Gueguen, C. & Tortell, P. D. 2008. High-resolution measurement of
Southern Ocean CO2 and O2 ⁄ Ar by membrane inlet mass
spectrometry. Mar. Chem. 108:184–94.

Hall-Spencer, J. M., Rodolfo-Metalpa, R., Martin, S., Ransome, E.,
Fine, M. E., Turner, S. M., Rowley, S. J., Tedesco, D. & Buia,
M.-C. 2008. Volcanic carbon dioxide vents show ecosystem
effects of ocean acidification. Nature 454:96–9.

Hein, M. & Sand-Jensen, K. 1997. CO2 increases oceanic primary
productivity. Nature 388:526–7.

Hinga, K. R. 2002. Effects of high pH on coastal marine phyto-
plankton. Mar. Ecol. Prog. Ser. 238:281–300.

Hunter, K. A. 2007. SWCO2 software for computation of equilib-
rium composition of carbon dioxide in seawater. Available at:
http://neon.otago.ac.nz/research/mfc/people/keith_hunter/
software/software.htm.

Hurd, C. L. 2000. Water motion, marine macroalgal physiology,
and production. J. Phycol. 36:453–72.

Hurlbert, S. H. 1984. Pseudoreplication and the design of ecolog-
ical field experiments. Ecol. Monogr. 54:187–211.

Hutchins, D. A., Fu, F.-X., Zhang, Y., Warner, M. E., Feng, Y.,
Portune, K., Bernhardt, P. W. & Mulholland, M. R. 2007. CO2
control of Trichodesmium N2 fixation, photosynthesis, growth
rates, and elemental ratios: implications for past, present and
future ocean biogeochemistry. Limnol. Oceanogr. 52:1293–304.

Iglesias-Rodriguez, M. D., Buitenhuis, E. T., Raven, J. A., Schofield,
O., Poulton, A. J., Gibbs, S., Halloran, P. R. & de Baar, H. J. W.
2008b. Response to comment on ‘‘Phytoplankton calcification
in a high-CO2 world.’’ Science 322:1466c.

Iglesias-Rodriguez, M. D., Halloran, P. R., Rickaby, R. E. M., Hall, I.
R., Colmenero-Hidalgo, E., Gittins, J. R., Green, D. R. H., et al.

2008a. Phytoplankton photosynthesis in a high-CO2 world.
Science 320:336–40.

Israel, A. & Hophy, M. 2002. Growth, photosynthetic properties and
Rubisco activities and amounts of marine macroalgae grown
under current and elevated seawater CO2 concentrations. Glob.
Change Biol. 8:831–40.

Israel, A., Katz, S., Dubinsky, Z., Merrill, J. E. & Friedlander, M.
1999. Photosynthetic inorganic carbon utilization and
growth of Porphyra linearis (Rhodophyta). J. Appl. Phycol.
11:447–53.

Kübler, J. E., Johnston, A. M. & Raven, J. A. 1999. The effects of
reduced and elevated CO2 and O2 on the seaweed Lomentaria
articulata. Plant. Cell Environ. 22:1303–10.

Kuffner, I. B., Andersson, A. J., Jokiel, P. L., Rodgers, K. S. &
Mackenzie, F. 2007. Decreased abundance of crustose coral-
line algae due to ocean acidification. Nat. Geosci. 1:114–7.

Langdon, C. 2003. Review of experimental evidence for effects of
CO2 on calcification of reef builders. Glob. Biogeochem. Cycles
17:1011.

Langer, G., Geisen, M., Baumann, K.-H., Kläs, J., Riebesell, U.,
Thoms, S. & Young, J. R. 2006. Species-specific responses of
calcifying algae to changing seawater carbonate chemistry.
Geochem. Geophys. Geosyst. 7:1–12.

Leclercq, N., Gattuso, J.-P. & Jaubert, J. 2000. CO2 partial pressure
controls the calcification rate of a coral community. Glob.
Change Biol. 6:329–34.

Leonardos, N. & Geider, R. J. 2005. Elevated atmospheric carbon
dioxide increases organic carbon fixation by Emiliania huxleyi
(Haptophyta), under nutrient-limited high-light conditions.
J. Phycol. 41:1196–203.

Levitan, O., Rosenberg, G., Setlik, I., Setilkova, E., Grigel, J.,
Klepetar, J., Prasilm, O. & Berman-Frank, I. 2007. Ele-
vated CO2 enhances nitrogen fixation and growth in the
marine cyanobacterium Trichodesmium. Glob. Change Biol.
13:531–8.

Lewis, E. & Wallace, D. W. R. 1998. Program Developed for CO2 System
Calculations. ORNL ⁄ CDIAC-105. Carbon Dioxide Information
Analysis Center, Oak Ridge National Laboratory, U.S.
Department of Energy, Oak Ridge, Tennessee.

Maberly, S. C. 1990. Exogenous Sources on inorganic carbon for
photosynthesis by marine macroalgae. J. Phycol. 26:439–49.

Mackenzie, F. T. & Lerman, A. 2006. Carbon in the Geobiosphere –
Earth’s Outer Shell. Springer, Dordrecht, the Netherlands, 402
pp.

McNeil, B. I. & Matear, R. J. 2008. Southern Ocean acidification: a
tipping point at 450-ppm atmospheric CO2. Proc. Natl. Acad.
Sci. U. S. A. 105:18860–4.

Midelboe, A. L. & Hansen, P. J. 2007. High pH in shallow-water
macroalgal habitats. Mar. Ecol. Prog. Ser. 338:107–17.

Nimer, N. A. & Merrett, M. J. 1993. Calcification rate in Emiliania
huxleyi Lohmann in response to light, nitrate and availability of
inorganic carbon. New Phytol. 123:673–7.

Ohline, S. M., Reid, M. R., Husheer, S. L. G., Currie, K. & Hunter,
K. A. 2007. Spectrophotometric determination of pH in sea-
water off Taiaroa Head, Otago, New Zealand: full-spectrum
modeling and accurate prediction of pCO2 levels. Mar. Chem.
107:143–55.

Orr, J. C., Fabry, V. J., Aumont, O., Bopp, L., Doney, S. C., Feely, R.
A., Gnanadesikan, A., et al. 2005. Anthropogenic ocean acid-
ification over the twenty-first century and its impact on calci-
fying organisms. Nature 437:681–6.

Parsons, R., Raven, J. A. & Sprent, J. I. 1992. A simple flow system
used to measure acetylene reduction activity of Sesbania rostrata
stem and root nodules. J. Exp. Bot. 43:595–604.

Pruder, G. D. & Bolton, E. T. 1980. Differences between cell divi-
sion and carbon fixation rates associated with light intensity
and oxygen concentration: implications in the cultivation of
an estuarine diatom. Mar. Biol. 59:1–6.

Raupach, M. R., Marland, G., Ciais, P., Le Quere, C., Canadell, J. G.,
Klepper, G. & Field, C. B. 2007. Global and regional drivers
of accelerating CO2 emissions. Proc. Natl. Acad. Sci. U. S. A.
104:10288–93.

1250 C A T R I O N A L . H U R D E T A L .

Raven, J. A., Ball, L. A., Beardall, J., Giordano, M. & Maberly, S. C.
2005. Algae lacking carbon concentrating mechanisms. Can. J.
Bot. 83:879–90.

Raven, J. A. & Geider, R. D. 2003. Adaptation, acclimation and
regulation of photosynthesis in algae. In Larkum, A. W. D.,
Douglas, S. E. & Raven, J. A. [Eds.] Photosynthesis in Algae.
Kluwer Academic Publishers, Dordrecht, the Netherlands,
pp. 385–412.

Raven, J. A., Kübler, J. & Beardall, J. 2000. Put out the light, and
then put out the light. J. Mar. Biol. Assoc. U. K. 80:1–25.

Riebesell, U., Bellerby, R. G. J., Engel, A., Fabry, V. J., Hutchins, D.
A., Reusch, T. B. H., Shulz, K. G. & Morel, F. M. M. 2008.
Comment on ‘‘Phytoplankton calcification in a high-CO2
world.’’ Science 322:1466b.

Riebesell, U., Zondervan, I., Rost, B., Tortell, P. D., Zeebe, R. E. &
Morel, F. M. M. 2000. Reduced calcification on marine
plankton in response to increased atmospheric CO2. Nature
407:364–7.

Roberts, R. A. 2001. Review of settlement cues for larval abalone
(Haliotis spp.). J. Shellfish Res. 20:571–86.

Schippers, P., Lürling, M. & Scheffer, M. 2004. Increased atmo-
spheric CO2 promotes phytoplankton productivity. Ecol. Lett.
7:446–51.

Shirayama, Y. & Thornton, H. 2005. Effect of increased atmo-
spheric CO2 on shallow water marine benthos. J. Geophys. Res.
Oceans 110:C09S08.

Smith, S. V. 1981. Marine macrophytes as a global carbon sink.
Science 211:838–40.

Smith, A. D. & Roth, A. A. 1979. Effect of carbon dioxide con-
centration on calcification in the red coralline alga Bossiella
orbigniana. Mar. Biol. 52:217–25.

Solomon, S., Qin, D., Manning, M., Chen, Z., Marquis, M., Averyt,
K. B., Tignor, M. & Miller, H. L. [Eds.] 2007. Climate Change
2007: The Physical Science Basis. Contribution of Working Group I to
the Fourth Assessment Report of the Intergovernmental Panel on Cli-
mate Change. Cambridge University Press, Cambridge, UK, 996
pp.

Swanson, A. K. & Fox, C. H. 2007. Altered kelp (Laminariales)
phlorotannins and growth under elevated carbon dioxide and
ultraviolet-B treatments can influence associated intertidal
food webs. Glob. Change Biol. 13:1696–709.

Tapp, M., Hunter, K. A., Currie, K. & Macaskill, J. 2000. Apparatus
for continuous-flow underway spectrophotometric measure-
ment of surface water pH. Mar. Chem. 72:193–202.

The Royal Society. 2005. Ocean Acidification Due to Increasing Atmo-
spheric Carbon Dioxide. Policy Document 12 ⁄ 05. The Royal
Society, London, 57 pp.

Thomas, W. H. & Gibson, C. H. 1990. The effects of small-scale
turbulence on microalgae. J. Appl. Phycol. 2:71–7.

Tortell, P. D., DiTullio, G. R., Sigman, D. M. & Morel, F. M. M.
2002. CO2 effects on taxonomic composition and nutrient
utilization in an equatorial Pacific phytoplankton assemblage.
Mar. Ecol. Prog. Ser. 236:37–43.

Tortell, P. D., Payne, C. D., Li, Y., Trimborn, S., Rost, B., Smith, W. O.,
Riesselman, C., Dunbar, R. B., Sedwick, P. & DiTullio, G. R.
2008. CO2 sensitivity of Southern Ocean phytoplankton.
Geophys. Res. Lett. 35:L04605.

Zondervan, I. 2007. The effect of light, macronutrients, trace
metals and CO2 on the production of calcium carbonate and
organic carbon in coccolithophores – a review. Deep-Sea Res.
Part II Top. Stud. Oceanogr. 54:521–37.

O C E A N A C I D I F I C A T I O N A N D A L G A L M E T A B O L I S M 1251

Ocean Acidification and Its Potential Effects
on Marine Ecosystems

John M. Guinottea and Victoria J. Fabryb

aMarine Conservation Biology Institute, Bellevue, Washington, USA
bCalifornia State University San Marcos, San Marcos, California, USA

Ocean acidification is rapidly changing the carbonate system of the world oceans.
Past mass extinction events have been linked to ocean acidification, and the current
rate of change in seawater chemistry is unprecedented. Evidence suggests that these
changes will have significant consequences for marine taxa, particularly those that
build skeletons, shells, and tests of biogenic calcium carbonate. Potential changes in
species distributions and abundances could propagate through multiple trophic levels
of marine food webs, though research into the long-term ecosystem impacts of ocean
acidification is in its infancy. This review attempts to provide a general synthesis of
known and/or hypothesized biological and ecosystem responses to increasing ocean
acidification. Marine taxa covered in this review include tropical reef-building corals,
cold-water corals, crustose coralline algae, Halimeda, benthic mollusks, echinoderms,
coccolithophores, foraminifera, pteropods, seagrasses, jellyfishes, and fishes. The risk
of irreversible ecosystem changes due to ocean acidification should enlighten the ongo-
ing CO2 emissions debate and make it clear that the human dependence on fossil fuels
must end quickly. Political will and significant large-scale investment in clean-energy
technologies are essential if we are to avoid the most damaging effects of human-induced
climate change, including ocean acidification.

Key words: ocean acidification; climate change; carbonate saturation state; seawater
chemistry; marine ecosystems; anthropogenic CO2

Introduction

The carbonate system (pCO2, pH, alkalin-
ity, and calcium carbonate saturation state) of
the world oceans is changing rapidly due to
an influx of anthropogenic CO2 (Skirrow &
Whitfield 1975; Whitfield 1975; Broecker &
Takahashi 1977; Broecker et al. 1979; Feely
& Chen 1982; Feely et al. 1984; Kleypas et al.
1999a; Caldeira & Wickett 2003; Feely et al.
2004; Orr et al. 2005). Ocean acidification
may be defined as the change in ocean chem-
istry driven by the oceanic uptake of chemi-
cal inputs to the atmosphere, including carbon,
nitrogen, and sulfur compounds. Today, the

Address for correspondence: John M. Guinotte, Marine Conserva-
tion Biology Institute, 2122 112th Avenue NE, Suite B-300, Belle-
vue, WA 98004-2947. Voice: +1-425-274-1180; fax: +1-425-274-1183.
john@mcbi.org

overwhelming cause of ocean acidification is
anthropogenic atmospheric CO2, although in
some coastal regions, nitrogen and sulfur are
also important (Doney et al. 2007). For the past
200 years, the rapid increase in anthropogenic
atmospheric CO2, which directly leads to de-
creasing ocean pH through air–sea gas ex-
change, has been and continues to be caused
by the burning of fossil fuels, deforestation, in-
dustrialization, cement production, and other
land-use changes. The current rate at which
ocean acidification is occurring will likely have
profound biological consequences for ocean
ecosystems within the coming decades and
centuries.

Presently, atmospheric CO2 concentration is
approximately 383 parts per million by volume
(ppmv), a level not seen in at least 650,000 years,
and it is projected to increase by 0.5% per year

Ann. N.Y. Acad. Sci. 1134: 320–342 (2008). C© 2008 New York Academy of Sciences.
doi: 10.1196/annals.1439.013 320

Guinotte & Fabry: Ocean Acidification and Marine Ecosystems 321

TABLE 1. Projected changes in surface ocean carbonate chemistry based on IPCC IS92a CO2 emission
scenario (Houghton et al. 2001)a

Parameter Symbol Unit Glacial Preindustrial Present 2 × CO2 3 × CO2
Temperature T ◦C 15.7 19 19.7 20.7 22.7
Salinity S 35.5 34.5 34.5 34.5 34.5
Total alkalinity AT µmol kg−1 2356 2287 2287 2287 2287
pCO2 in seawater pCO2 µatm 180 280 380 560 840

(−56) (0) (35.7) (100) (200)
Carbonic acid H2CO3 µmol kg−1 7 9 13 18 25

(−29) (0) (44) (100) (178)
Bicarbonate ion HCO3 − µmol kg−1 1666 1739 1827 1925 2004

(−4) (0) (5) (11) (15)
Carbonate ion CO3 2− µmol kg−1 279 222 186 146 115

(20) (0) −(16) (−34) (−48)
Hydrogen ion H+ µmol kg−1 4.79 × 10−3 6.92 × 10−3 8.92 × 10−3 1.23 × 10−2 1.74 × 10−2

(−45) (0) (29) (78) (151)
Calcite saturation �calc 6.63 5.32 4.46 3.52 2.77

(20) (0) (−16) (−34) (−48)
Aragonite saturation �arag 4.26 3.44 2.9 2.29 1.81

(19) (0) (−16) (−33) (−47)
Dissolved inorganic DIC µmol kg−1 1952 1970 2026 2090 2144

carbon
(−1) (0) (2.8) (6.1) (8.8)

Total pH pHT 8.32 8.16 8.05 7.91 7.76

aWe assume that PO4 = 0.5 µmol L−1 and Si = 4.8 µmol L−1, and use the carbonic acid dissociation constants of
Mehrbach et al. (1973) as refit by Dickson and Millero (1987). pHT is based on seawater scale. Percent change from
preindustrial values are in parentheses. After Feely et al. (2008).

throughout the 21st century (Petit et al. 1999;
Houghton et al. 2001; Augustin et al. 2004;
Siegenthaler et al. 2005; Meehl et al. 2007). The
rate of current and projected increases in atmo-
spheric CO2 is approximately 100× faster than
has occurred in at least 650,000 years (Siegen-
thaler et al. 2005). In recent decades, only half of
anthropogenic CO2 has remained in the atmo-
sphere; the other half has been taken up by the
terrestrial biosphere (ca. 20%) and the oceans
(ca. 30%) (Feely et al. 2004; Sabine et al. 2004).
Since the Industrial Revolution, a time span of
less than 250 years, the pH of surface oceans has
dropped by 0.1 pH units (representing an ap-
proximately 30% increase in hydrogen ion con-
centration relative to the preindustrial value)
and is projected to drop another 0.3–0.4 pH
units by the end of this century (Mehrbach et al.
1973; Lueker et al. 2000; Caldeira & Wickett

2003; Caldeira et al. 2007; Feely et al. 2008).
[Note: The pH scale is logarithmic, and as a
result, each whole unit decrease in pH is equal
to a 10-fold increase in acidity.] A pH change
of the magnitude projected by the end of this
century probably has not occurred for more
than 20 million years of Earth’s history (Feely
et al. 2004). The rate of this change is cause
for serious concern, as many marine organ-
isms, particularly those that calcify, may not be
able to adapt quickly enough to survive these
changes.

A series of chemical reactions is initiated
when CO2 is absorbed by seawater. � is the
calcium carbonate saturation state:

� = [Ca2+][CO2−3 ]/K∗sp
where K∗sp is the stoichiometric solubility

product for CaCO3 and [Ca2+] and [CO
2−
3 ]

322 Annals of the New York Academy of Sciences

TA
B
LE

2
.

C
h

a
n

g
e
s

in
ca

lc
ifi

ca
ti

o
n

in
re

sp
o
n
se

to
in

cr
e
a
se

d

p
C

O
2

co
n

ce

n
tr

a
ti
o
n
s

re
la

ti
ve

to
ca

lc
ifi
ca
ti
o
n

a
t

p
re

se
n
t-

d
a
y

p
C
O
2

(∼
3

8
0

p
p

m

v)
a

A
p

p
ro

x.
%

ch
an

g

e
in

ca
lc

ifi
ca

ti
on

w
h

en
p

C
O

2
is

O
rg

a

n
is

m
/S

ys
te

m
M

in
er

al
og

y
2×

p
re

in
d

u
st

ri
al

3×
p

re
in

d
u

st
ri

al
R

ef
er

en
ce

s
N

ot
es

C
o

cc
o

li
th

o
p

h
o

re
s

E
m

il
ia

n
ia

h
u
xl

ey
i

C
al

ci
te

−2
5

S
ci

an
d

r

a
et

a
l.

20
03

E
.
h
u
xl

ey
i

′′
−9

−1
8

R
ie

b
es

el
le

t
a
l.

20
00

;
D

ec
re

as
e

in
C

aC
O

3
/c

el
l

Z
on

d
er

va
n

et
a
l.

20
01

E
.
h
u
xl
ey
i

′′
−4

0
D

el
ill

e
et

a
l.

20
05

I

n
it

ia
lp

C
O

2
=

71
3

p
p

m
v.

D
ec

re
as

e
in

n
et

co
m

m
u

n
it

y
ca

lc
ifi
ca
ti

on
G

ep
h
yr

oc
a
p
sa

oc
ea

n
ic

a
′′

−2
9

−6
6

R
ie
b
es
el
le
t
a
l.
20
00
;
D
ec
re
as
e
in
C
aC
O
3
/c
el
l
Z
on
d
er
va
n
et
a
l.
20
01

C
oc

co
li
th

u
s

p
el

a
gi

cu
s

′′
0

0
L

an
ge

r
et

a
l.

20
06

N
o

ch
an

ge
in

C
aC

O
3
/c

el
l

C
a
lc

id
is

cu
s

le
pt

op
or

u
s
−2
5

L
an

ge
r

et
a
l.
20
06
D
ec
re
as
e
in
C
aC

O
3
;d

ec
re
as
e

al
so

ob
se

rv
ed

at
18

0
p

p
m

v
F

o
ra

m
in

if
er

a
O

rb
u
li
n
a

u
n
iv

er
sa

C
al
ci
te

−8
−1

4
S

p
er

o
et

a
l.

19
97

;
D
ec
re
as
e

in
sh

el
lw

ei
gh

t
B

ijm
a

et
a
l.

19
99

;
B

ijm
a
et
a
l.

20
02

G
lo

b
ig

er
in

oi
d
es

sa
cc

u
li
fe

r
′′

−4
to

−6
−6

to
−8

B
ijm

a
et
a
l.
19
99
;
D
ec
re
as
e
in
sh
el
lw
ei
gh
t
B
ijm
a
et
a
l.
20
02

S
cl

er
a

ct
in

ia
n

co
ra

ls
S
ty

lo
p
h
or

a
p
is

ti
ll
a
ta

′′
−1

4
−2

0
G

at
tu

so
et

a
l.

19
98

C
u

rv
e

w
as

n
on

lin
ea

r,
lin

ea
r

fit
w

as
m

a

d
e

fo
r
�

ar
ag

fr
om

2.
0−

3.
9

S
.
p
is

ti
ll
a
ta
′′
0

to
−5

0
R

ey
n

au
d

et
a
l.
20
03

L
ev

el
of

re
sp

on
se

is
te

m
p

er
at

u
re

–
d

ep
en

d
en

t
A

cr
op

or
a

ce
rv

ic
or

n
is

A
ra

go
n

it
e

−4
0

−5
9

R
en

eg
ar

&
R

ie
gl

20
05

Guinotte & Fabry: Ocean Acidification and Marine Ecosystems 323

TA
B
LE
2
.
A
p
p
ro
x.
%
ch
an
ge
in
ca
lc
ifi
ca
ti
on
w
h
en
p
C
O
2
is
O
rg

an
is

m
/S
ys
te
m
M
in
er
al
og
y
2×
p
re
in
d
u
st
ri
al
3×
p
re
in
d
u
st
ri
al
R
ef
er
en
ce
s
N
ot
es

A
cr

op
or

a
eu

ry
st

om
a

′′
−5

5
S

ch
n

ei
d

er
&

E
re

z
20

06
A

cr
op
or
a

ve
rw

ey
i
′′
−1

2
−1

8
M

ar
u

b
in

ie
t
a
l.

20
03

P.
co

m
p
re

ss
a

+

M
on

ti
p
or

a
ca

pi
ta

ta
′′

−4
0
−5
9
L
an

gd
on

&
A

tk
in

so
n

20
05

P
or

it
es

co
m
p
re
ss
a
′′
−1

7
−2

5
M

ar
u
b
in
ie
t
a
l.
20
01

P.
lu

te
a

′′
−3

8
−5

6
O

h
d

e
&

H
os

sa
in

20
04

P.
lu
te
a
′′
−3

3
−4

9
H

os
sa

in
&

O
h

d
e
20
06

P.
p
or

it
es
′′
−1

6
M

ar
u
b
in

i

&
T

h
ak

e
19

99
P

av
on

a
ca

ct
u
s

′′
−1
4
−2

0
M

ar
u
b
in
ie
t
a
l.
20
03

F
u
n
gi

a
sp

.
′′

−4
7

−6
9

H
os
sa
in
an
d
O
h
d
e
20
06

G
a
la

xe
a

fa
sc

ic

u
la

ri
s

′′
−1
2
−1
8
M
ar
u
b
in
ie
t
a
l.
20
03

G
.
fa

sc
ic

u
la
ri
s
′′
−5

6
−8

3
M

ar
sh

al
l&

C
lo

d
e
20
02

T
u
rb

in
a
ri

a
re

n
if
or

m
is

′′
−9

−1
3

M
ar

u
b

in
ie

t
a
l.
20
03
C
o

ra
ll

in
e

re
d

a
lg

a
e

P
or

ol
it
h
on

ga
rd

in
er

i
H

ig
h

-M
g

ca
lc
it
e
−2
5

A
ge

gi
an

19
85

M
es

o
co

s

m
s

a
n

d
fi

el
d

st
u

d
ie

s
B

io
sp

h
er

e
2

M
ix

ed
−5

6
−8

3
L

an
gd

on
et

a
l.
20
00

D
om

in
at

ed
by

co
ra

lli
n

e
re

d
al

ga
e

M
on

a
co

m
es

oc
os

m
′′

−2
1

L
ec

le
rc

q
et

a
l.
20
00
M
on
a
co
m
es
oc
os
m
′′

−1
5

L
ec
le
rc
q
et
a
l.
20
02

B
a
h
a
m

a
s

B
a
n
k

′′
−5

7
−8

5
B

ro
ec

ke
r

&
T

ak
ah

as
h

i1
96

6
R

u
ka

n
-s

h
o,

O
ki

n
a
w

a
′′

−4
5

−6
7

O
h
d
e

&
va

n
W

oe
si

k
19

99
N

a
tu

re
R

es
er

ve
R

ee
f,

R
ed

S
ea

′′
−5
5
S

ilv
er

m
an

et
a
l.

20
07

a
N

ea
rl

y
al

lo
rg

an
is
m
s

te
st

ed
to

d
at

e
sh

ow
re

d
u

ce
d

ca
lc
ifi
ca
ti
on
in
re

sp
on

se
to

el
ev

at
ed

p
C
O
2
an
d

d
ec

re
as

ed
ca

rb
on

at
e

io
n

co
n

ce
n

tr
at

io
n
an
d

ca
rb

on
at

e
sa

tu
ra

ti
on

st
at

e.
M

od
ifi

ed
fr

om
K

le
yp

as
et

a
l.

(2
00

6)
.

324 Annals of the New York Academy of Sciences

are the in situ calcium and carbonate concen-
trations, respectively. The end products of these
reactions are an increase in hydrogen ion con-
centration (H+), which lowers pH (making wa-
ters more acidic), and a reduction in the num-
ber of carbonate ions (CO2−3 ) available. This
reduction in carbonate ion concentration also
leads to a reduction in calcium carbonate sat-
uration state (�), which has significant impacts
on marine calcifiers. Table 1 lists carbon sys-
tem parameters and temperature changes for
surface waters based on the Intergovernmental
Panel on Climate Change (IPCC) IS92a CO2
emission scenario.

A reduction in the number of carbon-
ate ions available will make it more diffi-
cult and/or require marine calcifying organ-
isms to use more energy to form biogenic
calcium carbonate (CaCO3). Many marine
organisms form biogenic calcium carbonate
including: crustose coralline algae (the pri-
mary cementer that makes coral reef formation
possible), Halimeda (macroalgae), foraminifera,
coccolithophores, tropical reef-building corals,
cold-water corals, bryozoans, mollusks, and
echinoderms. The majority of marine calci-
fiers tested to date are sensitive to changes
in carbonate saturation state and have shown
declines in calcification rates in laboratory
and mesocosm studies (Table 2). These or-
ganisms are affected and will continue to be
affected by ocean acidification, but less well
known are the ecosystem impacts on higher
trophic-level organisms that rely on these cal-
cifiers for shelter, nutrition, and other core
functions.

Decreasing pH is not the only effect on the
inorganic carbon system in seawater that re-
sults from the ocean’s uptake of anthropogenic
CO2. Calcite and aragonite are the major
biogenically formed carbonate minerals pro-
duced by marine calcifiers, and the stability
of both minerals is affected by the amount
of CO2 in seawater, which is partially deter-
mined by temperature. Colder waters natu-
rally hold more CO2 and are more acidic than
warmer waters. The depths of the aragonite

and calcite saturation horizons are important
to marine calcifiers because the depth of these
horizons determines the limit at which pre-
cipitation of biogenic calcium carbonate by
marine organisms is favored (shallower than
the saturation horizon) and at which they will
experience dissolution (deeper than the satu-
ration horizons) in the absence of protective
mechanisms.

The aragonite and calcite saturation hori-
zons of the world’s oceans are moving to
shallower depths due to the rapid influx of an-
thropogenic CO2 to the oceans (Fig. 1). This
process has been well documented and mod-
eled at the global scale (Skirrow & Whitfield
1975; Broecker & Takahashi 1977; Feely &
Chen 1982; Feely et al. 1984, 1988; Kleypas
et al. 1999a; Broecker 2003; Caldeira & Wickett
2003; Feely et al. 2004; Caldeira & Wickett
2005; Orr et al. 2005). Future estimates of arag-
onite saturation horizon depth indicate that
shoaling will occur in the North Pacific, North
Atlantic, and Southern Ocean within the cen-
tury (Orr et al. 2005). The aragonite and cal-
cite saturation horizons in the North Pacific
are currently very shallow (Feely et al. 2004)
and are moving toward the surface at a rate of
1–2 m per year (R.A. Feely pers. comm. 2007).
Many of the areas where shoaling is predicted
to occur within the century are highly produc-
tive and home to many of the world’s most
important and economically lucrative commer-
cial fisheries.

It is clear that human-induced changes in
atmospheric CO2 concentrations are funda-
mentally altering ocean chemistry from the
shallowest waters to the darkest depths of the
deep sea. The chemistry of the oceans is ap-
proaching conditions not seen in many mil-
lions of years, and the rate at which this is
occurring is unprecedented (Caldeira & Wick-
ett 2003). Caldeira and Wickett (2003, p. 365)
state “Unabated CO2 emissions over the com-
ing centuries may produce changes in ocean
pH that are greater than any experienced
in the past 300 million years, with the pos-
sible exception of those resulting from rare,

Guinotte & Fabry: Ocean Acidification and Marine Ecosystems 325

Figure 1. (A) Depth of the aragonite saturation horizon (ASH), locations of deep-sea
bioherm-forming corals, and diversity contours for 706 species of azooxanthellate corals.
Projected ASH depth for year 1765 (preindustrial); pCO2 = 278 ppmv. Green triangles
are locations of six deep-sea, scleractinian, bioherm-forming coral species (Lophelia pertusa,
Madrepora oculata, Goniocorella dumosa, Oculina varicosa, Enallopsammia profunda, and
Solenosmilia variabilis). Numerals not falling on diversity contours indicate number of azoox-
anthellate coral species. Reprinted with permission from Guinotte et al. (2006). (B) Projected
ASH depth for year 2040; pCO2 = 513 ppmv. (C) Projected ASH depth for year 2099; pCO2
= 788 ppmv. Black areas appearing in the Southern Ocean and North Pacific indicate areas
where the ASH depth has reached the surface. (In color in Annals online.)

326 Annals of the New York Academy of Sciences

catastrophic events in Earth history” (Caldeira
and Rampino 1993; Beerling and Berner
2002). Recent evidence suggests ocean acid-
ification was a primary driver of past mass
extinctions and reef gaps, which are time
periods on the order of millions of years
that reefs have taken to recover from mass
extinctions (Stanley 2006; Veron 2008). Za-
chos and colleagues (2005) calculated that if
the entire fossil fuel reservoir (ca. 4500 GtC)
were combusted, the impacts on deep-sea pH
and biota would probably be similar to those
in the Paleocene–Eocene Thermal Maximum
(PETM), 55 million years ago. The PETM
likely caused a mass extinction of benthic
foraminifera (Zachos et al. 2005). Projected an-
thropogenic carbon inputs will occur within just
300 years, which is thought to be much faster
than the CO2 release during the PETM and too
rapid for dissolution of calcareous sediments to
neutralize anthropogenic CO2. Consequently,
the ocean acidification-induced impacts on sur-
face ocean pH and biota will probably be more
severe than during the PETM (Zachos et al.
2005).

While it is apparent changing seawater
chemistry will have serious consequences for
many marine calcifiers, the effects of ocean
acidification on noncalcifiers and the ecosys-
tem responses to these changes will be com-
plex and difficult to quantify. Assessing whether
ocean acidification is the primary driver of
a species’ population decline will be diffi-
cult due to the multitude of ongoing phys-
ical and chemical changes currently occur-
ring in the ocean. Ocean acidification is oc-
curring in synergy with significant ongoing
environmental changes (e.g., ocean temper-
ature increases), and these cumulative im-
pacts or interactive effects of multiple stressors
may have more significant consequences for
biota than any single stressor. Thus, research
into the synergistic effects of these changes
on marine organisms and the consequent
ecosystem responses is critical but still in its
infancy.

Calcification and Dissolution
Response

Hermatypic Corals (Zooxanthellate)

The calcification response of reef-building
corals to decreases in aragonite saturation state
has been well documented for a handful of se-
lect species. These experiments have been con-
ducted in laboratory tanks and mesocosms, but
to date have not been conducted in in situ field
experiments under “natural” conditions. Evi-
dence from species tested to date indicate that
the calcification rates of tropical reef-building
corals will be reduced by 20–60% at double
preindustrial CO2 concentrations (pCO2 ca.
560 ppmv) (Gattuso et al. 1998; Kleypas et al.
1999a; Langdon et al. 2000; Kleypas & Lang-
don 2002; Langdon et al. 2003; Reynaud et al.
2003; Langdon & Atkinson 2005; c.f. Royal
Society 2005; c.f. Kleypas et al. 2006) (see
Table 2). Figure 2 illustrates the projected re-
duction in surface-water aragonite saturation
state through the year 2069. A reduction in cal-
cification of this magnitude could fundamen-
tally alter the current structure and function of
coral-reef ecosystems, as their growth is depen-
dent on their ability to accrete at faster rates
than erosional processes can break them down.
Reef accretion will become increasingly more
critical in the coming decades, as global sea
levels rise and available light for photosynthe-
sis becomes a limiting factor for corals at the
deepest reaches of the photic zone.

A substantial decrease in the number of car-
bonate ions available in seawater will have se-
rious implications for coral calcification rates
and skeletal formation. Weaker coral skele-
tons will probably result from a reduction in
carbonate ions, enabling erosional processes
to occur at much faster rates than have oc-
curred in the past, and slower growth rates
may also reduce corals’ ability to compete for
space and light, though no studies have been
conducted to test this hypothesis (reviewed in
Kleypas et al. 2006). Biosphere II mesocosm ex-
periments suggest that net reef dissolution will

Guinotte & Fabry: Ocean Acidification and Marine Ecosystems 327

Figure 2. (A) Surface aragonite saturation state. Calculated preindustrial (1870) �arag values; pCO2
= 280 ppmv. Green dots represent present-day distribution of zooxanthellate coral reefs. Figure modified
from Guinotte et al. (2003). Figure legend classification from Kleypas et al. (1999b). (B) Surface aragonite
saturation state. Projected �arag values, 2060–2069; pCO2 = 517 ppmv. (In color in Annals online.)

outpace net reef calcification when carbonate
ion concentration decreases to about 150 to
110 µmole kg−1, a range that corresponds to
atmospheric CO2 concentrations of 560–840
ppmv (C. Langdon pers. comm. 2007). Hoegh-
Guldberg and colleagues (2007) stated that
aragonite saturation values will favor erosion
when the carbonate ion concentration ap-
proaches 200 µmole kg−1 (atmospheric CO2
concentration = 480 ppmv). The effects of a
reduction in calcification rates on recruitment,
settlement, and juvenile life stages of most ma-
rine calcifiers, including the majority of scler-
actinian corals, are not well known. However,

Edmunds (2007) documented a decline in the
growth rates of juvenile scleractinian corals in
the U.S. Virgin Islands and raised the possi-
bility that the effects of global climate change
(increased seawater temperatures and decreas-
ing aragonite saturation state) have already re-
duced the growth rate of juvenile corals.

Fine and Tchernov (2007a) reported two
species of scleractinian corals were able to sur-
vive corrosive water conditions (pH values of
7.3–7.6), which caused their skeletons to dis-
solve completely, leaving the coral polyps ex-
posed. When water chemistry returned to nor-
mal/ambient conditions, the coral polyps were

328 Annals of the New York Academy of Sciences

able to recalcify their skeletons without any ob-
vious detrimental effects. These findings shed
new light on the hypothesis that corals have a
means of alternating between soft bodies and
skeletal forms, which are absent from the fos-
sil record during reef gaps (Stanley & Fautin
2001; Medina et al. 2006; Stanley 2006; Fine
& Tchernov 2007a). Fine and Tchernov’s re-
sults offer some hope for the future of corals
in a high CO2 world, but caution should be
exercised as these manipulative experiments
did not include the effects of predation on the
“naked” coral polyps. Hard skeletons also pro-
vide another core function for coral polyps by
protecting them from periodic natural events
such as tsunamis and cyclones, which can cause
significant damage to coral colonies and reef
systems.

There is some discrepancy regarding the rep-
resentativeness of the coral species used in the
Fine and Tchernov calcification experiments.
Stanley (2007) stated the experiments may
not be representative of all coral species, par-
ticularly zooxanthellate reef-building species,
which might have responded quite differently
to the experiments because of the complex
nature of their photosymbiosis. This assertion
was challenged by Fine & Tchernov (2007b)
in the statement that the evolution and physi-
ology of the studied species are indistinguish-
able from tropical reef-building species. Rep-
resentativeness aside, both parties agree that
ocean acidification poses a significant threat
to coral-reef ecosystems and the services they
provide.

Calcifying Macroalgae

Coralline Algae

Scleractinian corals are not the only reef-
calcifying organisms that are sensitive to de-
creasing saturation states. Crustose coralline al-
gae (CCA) are a critical player in the ecology of
coral-reef systems as they provide the “cement”
that helps stabilize reefs, make significant sed-
iment contributions to these systems, and are
important food sources for sea urchins, par-

rot fish, and several species of mollusks (Littler
& Littler 1984; Chisholm 2000; Diaz-Pulido
et al . 2007). CCA also provide important hard
settlement substrate for coral larvae (Heyward
& Negri 1999; Harrington et al. 2005; Diaz-
Pulido et al. 2007). Coralline algae produce cal-
cium carbonate in the form of high-magnesium
calcite, a more soluble form of calcium carbon-
ate than either calcite or aragonite, which make
these species particularly sensitive to decreasing
carbonate saturation states.

Mesocosm experiments exposing CCA to el-
evated pCO2 (2 × present day) indicate up to a
40% reduction in growth rates, 78% decrease
in recruitment, 92% reduction in total area cov-
ered by CCA, and a 52% increase in noncal-
cifying algae (Buddemeier 2007; Kuffner et al.
2008). Agegian (1985) also reported a reduc-
tion in recruitment when CCA were exposed to
elevated pCO2 in aquarium experiments. Bud-
demeier (2007) states, “The combined effects of
reduced carbonate production and diminished
stabilization (cementation) of coasts and shal-
low seafloors by encrusting calcifiers are likely
to lead to more rapid erosion and ecosystem
transitions (macroalgal takeover) than would
be expected on the basis of decreases in coral
growth alone.” The ecological importance of
coralline algae to reef systems and the effects
decreasing carbonate saturation state will have
on these organisms have been overlooked to a
significant degree, and more research is needed
to document CCA response to reduced car-
bonate saturation states and in turn how these
responses will impact reef ecosystems.

Halimeda

Halimeda is a genus of green, calcifying
macroalgae that forms extensive beds in cer-
tain regions of the world’s oceans. Some of
the most well-developed Halimeda beds occur
off the northeast coast of Australia, and es-
timates of total area covered by Halimeda in
the Great Barrier Reef region are upwards
of 2000 km2 (reviewed by Diaz-Pulido et al.
2007). Halimeda, along with other calcareous
algae (Udotea, Amphiroa, and Galaxaura), are

Guinotte & Fabry: Ocean Acidification and Marine Ecosystems 329

important producers of marine sediments
and contribute to reef accretion by filling
voids in the reef matrix with their sediments
(Hillis-Colinvaux 1980; Davies & Marshall
1985; Drew & Abel 1988; Diaz-Pulido et al.
2007). Reefs and Halimeda bioherms have high
calcification rates and are responsible for the
majority of CaCO3 production and accumu-
lation on the continental shelf (Milliman &
Droxler 1996; Kleypas et al. 2006).

The three-dimensional structures Halimeda
form, which can be 20 m in height, provide im-
portant habitat for adult fishes and may serve as
nursery grounds for juvenile fishes and inverte-
brates (Beck et al. 2003). Calcifying macroalgae
produce biogenic calcium carbonate in three
forms: high-magnesium calcite, aragonite, and
calcite; all of these forms are susceptible to the
negative effects of decreasing carbonate satura-
tion states (Littler & Littler 1984). Few species
of Halimeda have been exposed to high pCO2
in lab experiments, but one species from the
Great Barrier Reef, Halimeda tuna, displayed a
negative calcification response when exposed
to a pH drop of 0.5 units (8 to 7.5) (Borowitzka
& Larkum 1986).

Cold Water Corals (Azooxanthellate)

Cold water corals and the ecologically
rich bioherms they form are widely dis-
tributed throughout the world oceans (see
Fig. 1). A great number of these highly pro-
ductive ecosystems have been discovered only
in the last decade, and it is thought that the
area covered by these organisms may surpass
the total area of tropical zooxanthellate reef
systems (Mortensen et al. 2001; Freiwald et al.
2004; Freiwald & Roberts 2005; Guinotte et al.
2006; Turley et al. 2007). Cold water corals are
azooxanthellate, which means they do not con-
tain photosynthetic algae, and thus are not lim-
ited to the photic zone. The majority of cold
water corals are found in depths of 200–1000
m or more, and some solitary colonies have
been found at depths of several thousand me-
ters (Freiwald 2002; Freiwald et al. 2004). There

are six species of azooxanthellate, bioherm-
forming, scleractinian corals (Lophelia pertusa,
Madrepora oculata, Goniocorella dumosa, Oculina

varicosa, Enallopsammia profunda, and Solenosmilia
variabilis), all of which produce calcium carbon-
ate skeletons of aragonite. Cold water corals
bioherms have extremely high biodiversity and
provide habitat and nursery areas for many
deep-sea organisms, including several commer-
cially important fish species (Rogers 1999; Fossa
et al. 2002; Husebo et al. 2002). Scleractinian
cold water corals are not the only azooxanthel-
late habitat formers.

The “coral gardens” of the North Pacific are
biodiversity hotspots dominated by octocorals
(soft corals, stoloniferans, sea fans, gorgonians,
and sea pens) and stylasterids, the majority of
which produce calcite spicules and holdfasts
(Cairns & Macintyre 1992; Guinotte et al. 2006;
Stone 2006). Stone (2006) reported that 85%
of the economically important fish species ob-
served on submersible transects in waters off
the Aleutian Islands were associated with corals
and other emergent epifauna. The waters off
the Aleutian Islands have the highest abun-
dance and diversity of cold-water corals found
to date in high-latitude ecosystems (Heifetz et al.
2005; Stone 2006), but well-developed scler-
actinian bioherms are curiously absent from
this region even though scleractinian bioherm-
forming species are found in North Pacific wa-
ters (Guinotte et al. 2006).

The reason scleractinian bioherms are not
present in North Pacific waters could be a
function of the shallow depth of the arago-
nite saturation horizon and high dissolution
rates throughout the region (Guinotte et al.
2006). If this hypothesis is true, then decreas-
ing carbonate saturation state will probably
impact scleractinian cold-water corals earlier
than shallow-water reef builders. Cold-water
corals are bathed in cold, deep waters that
have naturally high levels of CO2 (global av-
erage �arag = 2). The low carbonate saturation
state environment in which they live probably
contributes to their slow growth/calcification
rates, which are an order of magnitude slower

330 Annals of the New York Academy of Sciences

than tropical zooxanthellate corals (global aver-
age �arag = 4). Indeed, some deeper cold-water
coral bioherms could already be experiencing
corrosive conditions with respect to aragonite
saturation state (�arag < 1), though no evidence of this has been documented.

Greater than 95% of the present day
distribution of bioherm-forming scleractinian
species occur in waters that are supersaturated
with aragonite (Guinotte et al. 2006). Future
aragonite saturation state projections from Orr
and co-authors (2005) indicate that 70% of scle-
ractinian cold-water coral bioherms could be
in undersaturated water with respect to arago-
nite by the end of the century (Guinotte et al.
2006; Turley et al. 2007) (see Fig. 1). Labora-
tory experiments are currently being conducted
to test whether cold water corals scleractinians
(Lophelia pertusa) are sensitive to decreasing arag-
onite saturation state (Riebesell pers. comm.),
but no lab experiments have been conducted
to test the sensitivity of cold-water octocorals
and stylasterids to decreasing carbonate satu-
ration states. Manipulative CO2 experiments
to determine cold-water coral sensitivity and
calcification response to decreasing carbonate
saturation states are a top priority for future
research (Guinotte et al. 2006; Kleypas et al.
2006; Roberts et al. 2006; Turley et al. 2007).

Benthic Mollusks, Bryozoans,
and Echinoderms

The physiological and ecological impacts of
increasing pCO2 on benthic mollusks, bry-
ozoans, and echinoderms are not well known,
and few manipulative experiments have been
carried out to determine sensitivity to ele-
vated pCO2 (Kleypas et al. 2006). The nega-
tive effects of acidic waters on bivalves have
been investigated in a small number of stud-
ies (Kuwatani & Nishii 1969; Bamber 1987,
1990; Michaelidis et al. 2005; Berge et al. 2006),
and only one investigated the negative calci-
fication response to pCO2 levels within the
range predicted by the IPCC (Gazeau et al.
2007). Gazeau and colleagues (2007) found

that calcification rates of the mussel (Mytilus
edulis) and Pacific oyster (Crassostrea gigas) can
be expected to decline linearly with increas-
ing pCO2, 25% and 10% respectively, by the
end of the century (ca. 740 ppmv, IPCC IS92a
scenario). Both species are important coastal
ecosystem engineers and represent a signif-
icant portion of global aquaculture produc-
tion (Gazeau et al. 2007). Bivalves that settle
in coastal estuarine areas may be particularly
vulnerable to anthropogenic ocean acidifica-
tion. These organisms naturally experience ex-
tremely high mortality rates (>98%) in their
transition from larvae to benthic juveniles (re-
viewed by Green et al. 2004), and any increase
in juvenile mortality due to ocean acidification
could have serious effects on estuarine bivalve
populations.

Kurihara and colleagues (2007) demon-
strated that increased pCO2 of seawater pro-
jected to occur by the year 2300 (pH 7.4) will
severely impact the early development of the
oyster Crassostrea gigas and highlighted the im-
portance of acidification effects on larval de-
velopment stages of marine calcifiers. Because
early life stages appear to be more sensitive
to environmental disturbance than adults and
most benthic calcifiers possess planktonic lar-
val stages, fluctuations in larval stages due to
high mortality rates may exert a strong influ-
ence on the population size of adults (Green
et al. 2004). Kurihara and co-authors (2004)
investigated the effects of increased pCO2 on
the fertilization rate and larval morphology of
two species of sea urchin embryos (Hemicentrotus
pulcherrimus and Echinometra mathaei) and found
the fertilization rate of both species declined
with increasing CO2 concentration. In addi-
tion, the size of pluteus larvae decreased with
increasing CO2 concentration and malformed
skeletogenesis was observed in larval stages of
both species. Kurihara and Shirayama (2004)
concluded that both decreasing pH and altered
carbonate chemistry affect early development
and life history of many marine organisms,
which will result in serious consequences for
marine ecosystems.

Guinotte & Fabry: Ocean Acidification and Marine Ecosystems 331

Experiments focusing on the direct effects
of increasing ocean acidification on marine
calcifiers have been the dominant activity to
date, but numerous and ecologically significant
indirect effects are probable. Bibby and col-
leagues (2007) documented interesting behav-
ioral, metabolic, and morphological responses
of the intertidal gastropod Littorina littorea to
acidified seawater (pH = 6.6). This marine snail
produced thicker shells when exposed to preda-
tion (crab) cues in control experiments, but this
defensive response was disrupted when pH was
decreased. The snails also displayed reduced
metabolic rates and an increase in avoidance
behavior, both of which could have significant
ecosystem implications via organism interac-
tions, energy requirements, and predator–prey
relationships. This study investigated only one
species of mollusk, but other marine organisms
will probably have indirect responses to ocean
acidification (Bibby et al. 2007).

Coccolithophores, Foraminifera,
and Pteropods

The major planktonic producers of CaCO3
are coccolithophores (single-celled algae),
foraminifera (protists), and euthecosomatous
pteropods (planktonic snails). Coccol-
ithophores and foraminifera secrete CaCO3 in
the form of calcite, whereas pteropods secrete
shells made of aragonite, which is about 50%
more soluble in seawater than calcite (Mucci
1983). These planktonic groups differ with
respect to their size, trophic level, generation
time, and other ecological attributes. High
quality, quantitative data on the latitudinal and
vertical distributions and abundances of these
calcareous taxa are lacking, and estimates of
their contributions to global calcification rates
are poorly constrained.

The calcification response of coccol-
ithophores, foraminifera, and pteropods to
ocean acidification has been investigated to
date in very few species. Most studies have in-
volved bloom-forming coccolithophores, and
these species (Emiliania huxleyi and Geophyro-

capsa oceanica) show decreased calcification
rates ranging from 25 to 66% when pCO2
is increased to 560–840 ppmv, respectively
(TABLE 2) in lab and mesocosm experiments. In
lab experiments with the coccolithophore Coc-
colithus pelagicus, however, Langer et al. (2006)
found that calcification did not change with in-
creased CO2. Moreover, there is evidence sug-
gesting that at least one coccolithophore species
may have the capacity to adapt to changing
pCO2 over long periods. Experimental manip-
ulations show that Calcidiscus leptoporus exhibits
highest calcification rates at present-day CO2
levels, with malformed coccoliths and cocco-
spheres at both lower and higher pCO2 (Langer
et al. 2006). Because no malformed coccol-
iths were observed in sediments from the Last
Glacial Maximum (when pCO2 levels were
about 200 ppmv), the authors concluded that
C. leptoporus has adapted to present-day CO2
levels.

In lab experiments with two species of plank-
tonic foraminifera, shell mass decreased as the
carbonate ion concentration of seawater de-
creased (Spero et al. 1997; Bijma et al. 1999,
2002). When grown in lab experiments in
seawater chemistry equivalent to pCO2 val-
ues of 560 and 740 ppmv, shell mass of the
foraminifera Orbulina universa and Globigerinoides
sacculifer declined by 4–8% and 6–14%, respec-
tively, compared to the shell mass secreted at
the preindustrial pCO2 value.

Data for a single species of shelled pteropods
suggest that net shell dissolution occurs in live
pteropods when the aragonite saturation is
forced to <1.0 (Orr et al. 2005; Fabry et al. 2008). When live pteropods (Clio pyramidata) were collected in the subarctic Pacific and ex- posed to a level of aragonite undersaturation similar to that projected for Southern Ocean surface waters by the year 2100 under the IS92a emissions scenario, shell dissolution occurred within 48 hours, even though animals were ac- tively swimming.

The response of planktonic calcifying or-
ganisms to elevated pCO2 may not be uni-
form among species or over time. To date,

332 Annals of the New York Academy of Sciences

published research indicates that most cal-
careous plankton show reduced calcification
in response to decreased carbonate ion con-
centrations; however, the limited number of
species investigated precludes identification of
widespread or general trends. All studies thus
far on the impacts of ocean acidification
on calcareous plankton have been short-term
experiments, ranging from hours to weeks.
Nothing is known about the long-term impacts
of elevated pCO2 on the reproduction, growth,
and survivorship of planktonic calcifying or-
ganisms or their ability to adapt to changing
seawater chemistry. Chronic exposure to in-
creased pCO2 may have complex effects on the
growth and reproductive success of calcareous
plankton or may induce adaptations that are
absent in short-term experiments. No studies
have investigated the possibility of differential
impacts with life stage or age of the organism.
Additional experimental evidence from plank-
tonic calcifiers is urgently needed if we are
to develop a predictive understanding of the
impacts of ocean acidification on planktonic
communities.

Physiological Reponses

Fishes

Elevated CO2 partial pressures (hyper-
capnia) will affect the physiology of water-
breathing animals by inducing acidosis in the
tissues and body fluids of marine organisms,
including fishes (Roos & Boron 1981; Portner
et al. 2004). pH, bicarbonate, and CO2 lev-
els within the organism are altered with long-
term effects on metabolic functions, growth,
and reproduction, all of which could be harm-
ful at population and species levels (Portner
et al. 2004). Short-term effects of elevated CO2
on fishes include alteration of the acid–base
status, respiration, blood circulation, and ner-
vous system functions, while long-term effects
include reduced growth rate and reproduc-
tion (Ishimatsu & Kita 1999; Ishimatsu et al.

2004, 2005). Most experiments undertaken to
date involved altering pH to levels consistent
with conditions that would be present if CO2
were to be directly injected to the seafloor
(pH ca. 5.8–6.2). These experiments have
shown adverse negative effects of acidified sea-
water on fish throughout their entire life cycle
(eggs, larvae, juveniles, and adults) (Kikkawa
et al. 2003, 2004; Ishimatsu et al. 2004; Portner
et al. 2004).

Fish in early developmental stages are more
sensitive to environmental change than adults
and a limited number of studies have shown
this to be true when fish eggs, larvae, and juve-
niles were exposed to elevated CO2 (McKim
1977; Kikkawa et al. 2003, 2004; Ishimatsu
et al. 2004). Ishimatsu and co-authors (2004)
state, “Even if the severity of environmental
hypercapnia due to CO2 sequestration is made
tolerable to adults, a gradual reduction in pop-
ulation size and changes in marine ecosystem
structures are unavoidable consequences when
young individuals cannot survive” (p. 732). The
long-term effects and adaptation potential of
fishes experiencing future pCO2 levels consis-
tent with IPCC scenarios are not known.

Photosynthetic Organisms

Phytoplankton and Cyanobacteria

Most species of marine phytoplankton have
carbon-concentrating mechanisms that accu-
mulate inorganic carbon either as CO2 or
HCO−3 or both (Giordano et al. 2005). Owing
in large part to their carbon-acquisition mech-
anisms and efficiencies, most marine phyto-
plankton tested to date in single-species lab ex-
periments or natural community-perturbation
experiments show either no change or small
increases (generally ≤ 10%) in photosynthetic
rates when grown under high pCO2 condi-
tions equivalent to ca. 760 micro atmosphere
(µatm) (Tortell et al. 1997; Hein & Sand-Jensen
1997; Burkhardt et al. 2001; Tortell and Morel
2002; Rost et al. 2003; Beardall & Raven 2004;

Guinotte & Fabry: Ocean Acidification and Marine Ecosystems 333

Schippers et al. 2004; Giordano et al. 2005;
Martin & Tortell 2006). Unlike other major
phytoplankton groups investigated thus far, the
coccolithophorid Emiliania huxleyi has low affin-
ity for inorganic carbon and could be carbon-
limited in today’s ocean (Rost & Riebesell
2004). Whether E. huxleyi will show increased
rates of photosynthesis with progressive oceanic
uptake of atmospheric CO2, however, may de-
pend on nutrient availability and light condi-
tions (Zondervan 2007). In a recent mesocosm
CO2 manipulation, study, Riebesell and col-
leagues (2007) reported that CO2 uptake by a
phytoplankton community (primarily diatoms
and coccolithophores) in experimental pCO2
treatments of 700 and 1050 µatm was 27% and
39% higher, respectively, relative to the pCO2
treatment of 350 µatm.

Ocean acidification will be accompanied
by climate warming in large expanses of the
oceans. Higher sea-surface temperatures in-
crease thermal stratification of the upper ocean,
thereby reducing the vertical mixing of nutri-
ents to surface waters, and have been linked to
observed decreases in phytoplankton biomass
and productivity, particularly at low and mid-
latitudes (Behrenfeld et al. 2006). In warm,
nutrient-poor tropical and subtropical regions,
however, continued ocean absorption of an-
thropogenic CO2 may enhance fixation of
atmospheric nitrogen and could lead to in-
creased total primary productivity. Nitrogen-
fixing cyanobacteria in the genus Trichodesmium,
which support a large portion of primary pro-
ductivity in such low-nutrient areas of the
world’s oceans, show increased rates of nitro-
gen and carbon fixation under elevated pCO2
(Hutchins et al. 2007; Barcelos e Ramos et al.
2007). At CO2 levels of 750 ppmv, Trichodesmium
increased N2 fixation rates by 35–100% and
CO2 fixation rates by 15–128%, relative to
present-day CO2 conditions (Hutchins et al.
2007).

In a review of coastal marine phytoplankton,
Hinga (2002) found that while some species
grow well at a wide range of pH, others have
growth rates that vary greatly over a 0.5 to

1.0 pH unit change. He concluded that small
changes in ambient seawater pH could affect
species growth rates, abundances, and succes-
sion in coastal phytoplankton communities. Eu-
trophication and ocean acidification may act
in concert to amplify the pH range found in
coastal habitats, which in turn could lead to
increased frequency of blooms of those species
with tolerance to extreme pH (cf. Hinga 2002).
In both coastal and open ocean environments,
ocean acidification could also affect primary
productivity through pH-dependent speciation
of nutrients and metals (Zeebe & Wolf-Gladrow
2001; Huesemann et al. 2002).

Seagrasses

Seagrasses represent one of the most bio-
logically rich and productive marine ecosys-
tems in the ocean. They create critical nursery
grounds for juvenile fishes and important habi-
tat for adult fishes, invertebrates, and mollusks.
Several higher order and endangered species
rely on seagrasses for a significant portion of
their diet (e.g., dugongs, manatees, and green
sea turtles). Seagrass ecosystems are a critical
component to maintaining the biological diver-
sity of the oceans and could be one of the few
ecosystems that stand to benefit from increas-
ing levels of CO2 in seawater. Seagrasses are
capable of dehydrating HCO−3 , but many ap-
pear to use CO2 (aq) for at least 50% of their
carbon requirements used for photosynthesis
(Palacios & Zimmerman 2007). Zimmerman
and colleagues (1997) found that short-term
(ca. 45 days) CO2 (aq) enrichment increased
photosynthetic rates and reduced light require-
ments for eelgrass (Zostera marina L) shoots in
laboratory experiments.

Longer-term (1 year) experiments expos-
ing Zostera marina L to CO2 (aq) concentra-
tions of 36–1123 µM (pH 7.75–6.2) conducted
by Palacios and Zimmerman (2007) resulted
in higher reproductive output, an increase in
below-ground biomass, and vegetative prolifer-
ation of new shoots when light was in abundant
supply. These findings suggest that as the CO2

334 Annals of the New York Academy of Sciences

content of the surface ocean rises, so too will
the productivity of seagrass meadows, which
in turn may positively influence invertebrate
and fish populations. This increase in produc-
tivity will probably be true for other seagrass
species as most appear to be photosynthetically
limited by the present-day availability of CO2
(Durako 1993; Invers et al. 2001; Palacios and
Zimmerman 2007). Palacios and Zimmerman
(2007) noted that a significant indirect effect
of increased eelgrass density could be an in-
crease in sediment retention, which could lead
to increased water clarity and an expansion in
the depth distribution of eelgrasses to deeper
waters.

Community Impacts

Seagrasses, Coral Reefs, and Fishes

Seagrass meadows and mangroves provide
important nursery areas for juvenile fishes,
many of which migrate to coral reefs as adults,
and enhance fish diversity and abundance on
coral reefs adjacent to these ecosystems (Pollard
1984; Parrish 1989; Beck et al. 2001; Sheridan
& Hays 2003; Mumby et al. 2004; Dorenbosch
et al. 2005). The net effect of increasing CO2 on
seagrass ecosystems will probably be increased
seagrass biomass and productivity, assuming
water quality and clarity (low suspended sed-
iment) are sufficient for photosynthesis to oc-
cur. Under these conditions, it is probable that
an increase in total seagrass area will lead to
more favorable habitat and conditions for asso-
ciated invertebrate and fish species. However,
the net effect of ocean acidification on coral reef
ecosystems will probably be negative as many
reef-building marine calcifiers will be heavily
impacted by the combined effects of increasing
sea-surface temperatures (coral bleaching) and
decreasing carbonate saturation states of sur-
face waters in the coming decades (Guinotte
et al. 2003; Buddemeier et al. 2004). The mag-
nitude of both ecosystem responses to ocean
acidification and other environmental changes
working in synergy is difficult to predict as are

the net effects on fish abundance and diver-
sity. Predicting the net effects on fish popula-
tions is further complicated by the plethora of
unknowns surrounding the long-term effects of
increasing CO2 on fish physiology, metabolism,
and probable range shifts due to ocean
warming.

Cold-water Corals and Fishes

The ecology and species relationships of
cold-water coral ecosystems are not as ad-
vanced as the state of knowledge for shallow-
water coral-reef systems, which is due in large
part to logistical challenges and the expense
of operating vessels and submersibles in the
deep sea. However, cold-water coral ecosys-
tems are thought to provide important habi-
tat, feeding grounds, and recruitment/nursery
functions for many deep-water species, includ-
ing several commercially important fish species
(Mortensen 2000; Fossa et al. 2002; Husebo et al.
2002; Roberts et al. 2006). Many of the species
relationships are thought to be facultative, but
nonetheless, high fish densities have been re-
ported for these structure-forming ecosystems
(Husebo et al. 2002; Costello et al. 2005; Stone
2006). Populations of grouper, snapper, and
amberjack use the Oculina varicosa reefs off
the Florida coast as feeding and spawning ar-
eas (Reed 2002), even though their numbers
have been dramatically reduced by commer-
cial and recreational fishing in recent decades
(Koenig et al. 2000). Large aggregations of red-
fish (Sebastes spp.), ling (Molva molva), and tusk
(Brosme brosme Ascanius) have been documented
in the Lophelia pertusa reefs of the North Atlantic
(Husebo et al. 2002), and strong fish–coral asso-
ciations exist in the cold-water coral ecosystems
of the North Pacific (Stone 2006).

Ocean acidification could have significant
indirect effects on fishes and other deep-
sea organisms that rely on cold-water coral
ecosystems for protection and nutritional re-
quirements. Roberts and Gage (2003) docu-
mented over 1300 species living on the Lophelia

Guinotte & Fabry: Ocean Acidification and Marine Ecosystems 335

pertusa reefs in the NE Atlantic. Future depth
projections for the aragonite saturation hori-
zons indicate 70% of cold-water scleractinians
will be in undersaturated waters by the end of
the century, and significant decreases in cal-
cification rate could occur well before corals
experience undersaturated conditions as arag-
onite saturation state decreases progressively
over time (Guinotte et al. 2006). Quantifying
the indirect impacts of ocean acidification on
coral-associated fishes is not possible due to
uncertainties surrounding facultative and ob-
ligate species relationships, but the net effects
are likely to be negative as cold-water coral
growth, distribution, and area decrease.

Plankton

If reduced calcification decreases a calci-
fying organism’s fitness or survivorship, then
some planktonic calcareous species may un-
dergo shifts in their distributions as the inor-
ganic carbon chemistry of seawater changes.
Calcifying species that are CO−2 sensitive could
potentially be replaced by noncalcifying species
and/or those species not sensitive to elevated
pCO2.

By 2100, surface waters of polar and sub-
polar regions are projected to become un-
dersaturated with respect to aragonite (Orr
et al. 2005). Pteropods are important com-
ponents of the plankton in high-latitude sys-
tems, with densities reaching thousands of in-
dividuals m−3 (e.g., Bathmann et al. 1991;
Pane et al. 2004). If pteropods require sea-
water that is supersaturated with respect to
aragonite, then their habitat would become in-
creasingly limited, first vertically in the water
column and then latitudinally, by the shoal-
ing of the aragonite saturation horizon over
the next century (Feely et al. 2004; Orr et al.
2005). If high-latitude surface waters do be-
come undersaturated with respect to arago-
nite, pteropods could eventually be eliminated
from such regions, with consequences to food-
web dynamics and other ecosystem processes
(Fabry et al. 2008). In the subarctic Pacific,

for example, pteropods can be important prey
for juvenile pink salmon (Oncorhynchus gobuscha),
as well as chum and sockeye salmon, pollock,
and other commercially important fishes (Ay-
din pers. comm.). Armstrong and co-authors
(2005) reported interannual variability in the
diet of juvenile pink salmon, with a single
species of pteropod (Limacina helicina) compris-
ing 15 to 63% by weight of pink salmon di-
ets during a 3-year study. Because Pacific pink
salmon have a short, 2-year life cycle, prey
quality and abundance during the salmon’s ju-
venile stage may strongly influence the pink
salmon’s adult population size and biomass
(Aydin et al. 2005).

Jellyfish blooms (scyphomedusae, hydrome-
dusae, and cubomedusae) have increased over
the last several decades (Purcell et al. 2007), but
it is too soon to determine whether such recent
jellyfish increases will persist or the populations
will fluctuate with climatic regime shifts, par-
ticularly those at decadal scales, as has been
observed previously (Purcell 2005). Attrill and
colleagues (2007) reported a significant corre-
lation of jellyfish frequency in the North Sea
from 1971 to 1995 with decreased pH (from
8.3 to 8.1) of surface waters. Although the
causative mechanism is not known, Attrill and
colleagues (2007) suggest that projected climate
change and declining ocean pH will increase
the frequency of jellyfish in the North Sea
over the next century. Jellyfish are both preda-
tors and potential competitors of fish and may
substantially affect pelagic and coastal ecosys-
tems (Purcell & Arai 2001; Purcell 2005). It is
important to resolve possible linkages between
jellyfish blooms and ocean acidification and de-
termine whether continued changes in the sea-
water inorganic carbon system will exacerbate
problematic increases in jellyfish that have been
associated with climate change, overfishing, eu-
trophication, and other factors (Purcell et al.
2007).

Planktonic ecosystems are complex nonlin-
ear systems, and the consequences of ocean
acidification on such ecosystems are largely un-
known. Substantial changes to species diversity

336 Annals of the New York Academy of Sciences

and abundances, food-web dynamics, and
other fundamental ecological processes could
occur; however, the interactions and feedbacks
among the effects of chronic, progressively in-
creasing ocean acidification and other environ-
mental variables are difficult to predict. Ecosys-
tem responses will also depend on the ability of
biota to adapt to seawater chemistry changes
that are occurring at rates they have not en-
countered in their recent evolutionary history
(Siegenthaler et al. 2005). Future progress will
likely require integrated approaches involving
manipulative experiments, field observations,
and models, particularly at regional scales.

Summary and Conclusions

The scientific knowledge base surround-
ing the biological effects of ocean acidifica-
tion is in its infancy and the long-term con-
sequences of changing seawater chemistry on
marine ecosystems can only be theorized. Most
is known about the calcification response for
shallow-water scleractinian corals. Some data
sets allow the identification of “tipping points”
or “thresholds” of seawater carbonate chem-
istry when ocean acidification will cause net
calcification rates to be less than net dissolu-
tion rates in coral reef systems (Yates & Halley
2006; Hoegh-Guldberg et al. 2007). In contrast,
the potential effects ocean acidification may
have for the vast majority of marine species
are not known. Research into the synergistic
effects of ocean acidification and other human-
induced environmental changes (e.g., increas-
ing sea temperatures) on marine food webs
and the potential transformative effects these
changes could have on marine ecosystems is
urgently needed. It is important to have a firm
understanding of the degree to which ocean
acidification influences critical physiological
processes such as respiration, photosynthesis,
and nutrient dynamics, as these processes are
important drivers of calcification, ecosystem
structure, biodiversity, and ultimately ecosys-
tem health.

Future ocean acidification research needs in-
clude increased resources and efforts devoted
to lab, mesocosm, and in situ experiments, all
of which will aid in determining the biological
responses of marine taxa to increased pCO2.
Mesocosm and in situ experiments may simu-
late and/or provide more natural conditions
than single-species lab experiments, but they
have thus far used abrupt changes in seawater
chemistry which do not allow for potential ac-
climation or adaptation by marine organisms.
There is an additional need for experiments
on taxa with no commercial value but which
provide critical habitat and occupy impor-
tant trophic levels within marine food webs.
Direct CO2 experiments on commercially im-
portant species are clearly necessary, but non-
commercial species play crucial roles in marine
ecosystems and the life history of most com-
mercial species. The effects of ocean acid-
ification on less charismatic species and/or
species with no economic value should not
be overlooked. The biological response of ma-
rine organisms (both commercial and noncom-
mercial) to ocean acidification will be key to
making informed policy decisions that con-
form to sound ecosystem-based management
principles.

There is a critical need for well-developed
spatial and temporal models that give accu-
rate present day and future estimates of arago-
nite and calcite saturation states in the coastal
zones. The shallow continental shelves are
some of the most biologically productive ar-
eas in the sea and are home to the majority
of the world’s fisheries, but accurate carbonate
saturation state data do not currently exist for
most coastal regions. Ocean acidification in-
formation should also be integrated into exist-
ing ecosystem models, which attempt to predict
the effects of environmental changes on ma-
rine populations and ecosystem structure (e.g.,
Ecopath and Ecosim). Development of these
tools is essential to making credible predictions
of future ocean acidification effects on marine
ecosystems and will aid in guiding management
decisions.

Guinotte & Fabry: Ocean Acidification and Marine Ecosystems 337

The overwhelming volume of scientific evi-
dence collated by the IPCC documenting the
dangers of human-induced climate change, of
which ocean acidification is only one, should
end the lingering CO2 emissions reduction de-
bate. The global CO2 experiment which has
been under way since the Industrial Revolution
and the potentially dire consequences this
uncontrolled experiment poses for marine or-
ganisms and indeed, all life on Earth, leave no
doubt that human dependence on fossil fuels
must end as soon as possible. International col-
laboration, political will, and large-scale invest-
ment in clean energy technologies are essen-
tial to avoiding the most damaging effects of
human-induced climate change.

Acknowledgments

This work was supported in part by MCBI
grants from the Edwards Mother Earth Foun-
dation, Marisla Foundation, Moore Family
Foundation, and Mark and Sharon Bloome.
Support for VJF was provided in part by
National Science Foundation grants OCE-
0551726 and ANT-0538710. We would like
to thank RW Buddemeier, RA Feely, and an
anonymous reviewer for constructive inputs on
an early draft.

Conflict of Interest

The authors declare no conflicts of interest.

References

Agegian, C.R. 1985. The Biogeochemical Ecology of
Porolithon gardineri (Foslie). Ph.D. thesis, University
of Hawaii, Honolulu.

Armstrong, J.L. et al. 2005. Distribution, size, and interan-
nual, seasonal and diel food habits of northern Gulf
of Alaska juvenile pink salmon, Oncorhyncus gorbuscha.
Deep-Sea Res. II 52: 247–265.

Attrill, M., J. Wright & M. Edwards. 2007. Climate-
related increases in jellyfish frequency suggest a more
gelatinous future for the North Sea. Limnol. Oceanogr.
52: 480–485.

Augustin, L. et al. 2004. Eight glacial cycles from an
Antarctic ice core. Nature 429: 623–628.

Aydin, K.Y. et al. Linking oceanic foodwebs to coastal
production and growth rates to Pacific salmon (On-
corhynchus spp.), using models on three scales. Deep Sea
Res. II 52: 757–780.

Bamber, R.N. 1990. The effects of acidic seawater on 3
species of lamellibranch mollusk. J. Exp. Mar. Biol.
Ecol. 143: 181–191.

Bamber, R.N. 1987. The effects of acidic sea water on
young carpet-shell clams Venerupis decussata (L.)
(Mollusca: Veneracea). J. Exp. Mar. Biol. Ecol. 108:
241–260.

Barcelos e Ramos, J. et al. 2007. Effect of rising atmo-
spheric carbon dioxide on the marine nitrogen fixer
Trichodesmium. Global Biogeochem. Cycles 21: np.

Bathmann, U. et al. 1991. Short-term variations in par-
ticulate matter sedimentation off Kapp Norvegia,
Weddell Sea, Antarctica: relation to water mass ad-
vection, ice cover, plankton biomass and feeding ac-
tivity. Polar Biol. 11: 185–195.

Beardall, J. & J.A. Raven. 2004. The potential effects of
global climate change in microalgal photosynthesis,
growth and ecology. Phycologia 43: 31–45.

Beck, M.W. et al. 2001. The identification, conservation,
and management of estuarine and marine nurseries
for fish and invertebrates. Bioscience 51: 633–641.

Beck, M.W. et al. 2003. The role of nearshore ecosystems
as fish and shellfish nurseries. Issues Ecol. 11: 1–12.

Beerling, D.J. & R.A. Berner. 2002. Biogeochemical con-
straints on the Triassic-Jurassic boundary carbon cy-
cle event. Global Biogeochem. Cycles 16: 101–113.

Behrenfeld, M.J. et al. 2006. Climate-driven trends in
contemporary ocean productivity. Nature 444: 752–
755.

Berge, J.A. et al. 2006. Effects of increased sea water con-
centrations of CO2 on growth of the bivalve Mytilus
edulis L. Chemosphere 62: 681–687.

Bibby, R. et al. 2007. Ocean acidification disrupts induced
defences in the intertidal gastropod Littorina littorea.
Biol. Lett. 3: 699–701.

Bijma, J., H.J. Spero, & D.W. Lea. 1999. Reassessing
foraminiferal stable isotope geochemistry: Impact of
the oceanic carbonate system (experimental results).
In Use of Proxies in Paleoceanography: Examples from the
South Atlantic. G. Fischer & G. Wefer, Eds.: 489–512.
Springer-Verlag. New York.

Bijma, J., B. Honisch, & R.E. Zeebe. 2002. The im-
pact of the ocean carbonate chemistry on living
foraminiferal shell weight: Comment on “Carbon-
ate ion concentration in glacial-age deep waters of
the Caribbean Sea” by W.S. Broecker & E. Clark.
Geochem. Geophys. Geosyst. 3: 1064.

Borowitzka, L.J. & A.W.D. Larkum 1986. Reef algae.
Oceanus 29: 49–54.

338 Annals of the New York Academy of Sciences

Broecker, W.S. 2003. The Ocean CaCO3 Cycle. In The Oceans
and Marine Geochemistry. H. Elderfield, Ed.: 1–21.
Treatise on Geochemistry, Elsevier Pergamon. Ox-
ford.

Broecker, W.S. et al. 1979. Fate of fossil fuel carbon dioxide
and the global carbon budget. Science 206: 409–418.

Broecker, W.S. & T. Takahashi. 1966. Calcium carbonate
precipitation on the Bahama Banks. J. Geophys. Res.
71: 1575–1602.

Broecker, W.S. & T. Takahashi. 1977. Neutralization of
fossil fuel CO2 by marine calcium carbonate. In The
Fate of Fossil Fuel in the Oceans. N.R. Andersen & A.
Malahoff, Eds.: 213–241. Plenum Press. New York.

Buddemeier, R.W., J.A. Kleypas & R. Aronson. 2004.
Coral Reefs and Global Climate Change. Potential
Contributions of Climate Change to Stresses on
Coral Reef Ecosystems. Pew Center for Global Climate
Change. Arlington, VA. 42 pp.

Buddemeier, R.W. 2007. The future of tropical reefs and coast-
lines. Presented at the American Association for the Advance-

ment of Science Annual Meeting. San Francisco, CA, Feb
16.

Burkhardt, S. et al. 2001. CO2 and HCO3 uptake in ma-
rine diatoms acclimated to different CO2 concentra-
tions. Limnol. Oceanogr. 46: 1378–1391.

Cairns, S.D. & I.G. Macintyre. 1992. Phylogenetic im-
plications of calcium carbonate mineralogy in the
Stylasteridae (Cnidaria:Hydrozoa). Palaios 7: 96–
107.

Caldeira, K. & M.R. Rampino. 1993. Aftermath of the
end-Cretaceous mass extinction: possible biogeo-
chemical stabilization of the carbon cycle and cli-
mate. Paleoceanography 8: 515–525.

Caldeira, K. & M.E. Wickett. 2003. Anthropogenic car-
bon and ocean pH. Nature 425: 365.

Caldeira, K. & M.E. Wickett. 2005. Ocean model pre-
dictions of chemistry changes from carbon dioxide
emissions to the atmosphere and ocean. J. Geophys.
Res. 110: np.

Caldeira, K. et al. 2007. Comment on “Modern-
age buildup of CO2 and its effects on seawa-
ter acidity and salinity” by Hugo A. Loaiciga.
Geophysical Research Letters 34: L18608. doi:
10.1029/2006GL027288.

Chisholm, J.R.M. 2000. Calcification by crustose
coralline algae on the northern Great Barrier Reef,
Australia. Limnol. Oceanogr. 45: 1476–1484.

Costello, M. et al. 2005. Role of cold-water Lophelia
pertusa coral reefs as fish habitat in the NE At-
lantic. In Cold-water Corals and Ecosystems. A. Freiwald
& J.M. Roberts, Eds.: 771–805. Springer-Verlag.
Berlin, Heidelberg.

Davies, P.J. & J.F. Marshall. 1985. Halimeda bioherms – low
energy reefs, northern Great Barrier Reef. Proceedings
of the Fifth International Coral Reef Congress 1: 1–7.

Delille, B. et al. 2005. Response of primary production
and calcification to changes of pCO2 during exper-
imental blooms of the coccolithophorid Emiliania
huxleyi. Global Biogeochem. Cycles 19: np.

Diaz-Pulido, G. et al. 2007. Vulnerability of macroalgae of
the Great Barrier Reef to climate change. In Climate
Change and the Great Barrier Reef . J.E. Johnson & P.A.
Marshall, Eds.: 154–192. Great Barrier Reef Marine
Park Authority and Australian Greenhouse Office,
Australia.

Dickson, A.G. & F.J. Millero. 1987. A comparison of the
equilibrium constants for the dissociation of carbonic
acid in seawater media. Deep-Sea Res. A 34: 1733–
1743.

Doney, S.C. et al. 2007. The impacts of anthropogenic
nitrogen and sulfur deposition on ocean acidification
and the inorganic carbon system. Proc. Natl. Acad. Sci.
104: 14580–14585.

Dorenbosch, M. et al. 2005. Indo-Pacific seagrass beds and
mangroves contribute to fish density and diversity on
adjacent coral reefs. Mar. Ecol. Prog. Ser. 302: 63–76.

Drew, E.A., & K.M. Abel. 1988. Studies on Halimeda
I. The distribution and species composition of Hal-
imeda meadows throughout the Great Barrier Reef
Province. Coral Reefs 6: 195–205.

Durako, M.J. 1993. Photosynthetic utilization of CO2 (aq)
and HCO3 in Thalassia testudinum (Hydrocharita-
cae). Mar. Biol. 115: 373–380.

Edmunds, P.J. 2007. Evidence for a decadal-scale decline
in the growth rates of juvenile scleractinian corals.
Mar. Ecol. Prog. Ser. 341: 1–13.

Fabry, V.J. et al. 2008. Impacts of ocean acidification on
marine fauna and ecosystems processes. J. Mar. Sci.
In Press.

Feely, R.A. et al. 1988. Winter-summer variations of calcite
and aragonite saturation in the northeast Pacific.
Mar. Chem. 25: 227–241.

Feely, R.A., & C.T.A. Chen. 1982. The effect of excess
CO2 on the calculated calcite and aragonite satura-
tion horizons in the Northeast Pacific. Geophys. Res.
Lett. 9: 1294–1297.

Feely, R.A. et al. 2008. Present and future changes in
seawater chemistry due to ocean acidification. AGU
Monograph, The Science and Technology of CO2 Seques-
tration. In Press.

Feely, R.A. et al. 2004. Impact of anthropogenic CO2
on the CaCO3 system in the oceans. Science 305:
362–366.

Feely, R.A. et al. 1984. Factors influencing the degree of
saturation of the surface and intermediate waters of
the North Pacific Ocean with respect to aragonite.
J. Geophys. Res. 89: 631–640.

Fine, M. & D. Tchernov. 2007a. Scleractinian coral
species survive and recover from decalcification.
Science 315: 1811.

Guinotte & Fabry: Ocean Acidification and Marine Ecosystems 339

Fine, M. & D. Tchernov. 2007b. Response to “Ocean
acidification and scleractinian corals”. Science 317:
1032–1033.

Fossa, J.H., P.B. Mortensen, & D.M. Furevik. 2002. The
deep-water coral Lophelia pertusa in Norwegian wa-
ters: distribution and fishery impacts. Hydrobiologia
13: 1–12.

Freiwald, A. 2002. Reef-forming cold-water corals. In
Ocean Margin Systems. G. Wefer, D. Billett, D.B.B.
Hebbeln Jorgensen, et al., Eds.: 365–385. Springer.
Heidelberg.

Freiwald, A. et al. 2004. Coldwater Coral Reefs. UNEP-
WCMC, Cambridge, UK.

Freiwald, A. & J.M. Roberts. 2005. Cold-Water Corals and
Ecosystems. Springer. Heidelberg.

Gattuso, J.-P. et al. 1998. Effect of calcium carbonate satu-
ration of seawater on coral calcification. Global Planet.
Change 18: 37–46.

Gazeau, F. et al. 2007. Impact of elevated CO2 on
shellfish calcification. Geophys. Res. Lett. L07603,
doi:10.1029/2006GL028554.

Giordano, M., J. Beardall & J.A. Raven. 2005. CO2 con-
centrating mechanisms in algae: mechanisms, en-
vironmental modulation, and evolution. Annu. Rev.
Plant Biol. 56: 99–131.

Green, M.A. et al. 2004. Dissolution mortality of juvenile
bivalves in coastal marine deposits. Limnol. Oceanogr.
49: 727–734.

Guinotte, J.M., R.W. Buddemeier, & J.A. Kleypas. 2003.
Future coral reef habitat marginality: temporal and
spatial effects of climate change in the Pacific basin.
Coral Reefs 22: 551–58.

Guinotte, J.M. et al. 2006. Will human induced changes
in seawater chemistry alter the distribution of deep-
sea scleractinian corals? Front. Ecol. Environ. 4: 141–
146.

Harrington, L. et al. 2005. Synergistic effects of diuron
and sedimentation on photosynthesis and survival of
crustose coralline algae. Mar. Pollut. Bull. 51: 415–
427.

Heifetz, J. et al. 2005. Corals of the Aleutian Islands. Fish-
eries Oceanogr. 14: 131–138.

Hein, M. & K. Sand-Jensen. 1997. CO2 increases oceanic
primary production. Nature 388: 526.

Heyward, A.J. & A.P. Negri. 1999. Natural inducers of
coral larval metamorphosis. Coral Reefs 18: 273–
279.

Hillis-Colinvaux, L. 1980. Ecology and taxonomy of Hal-
imeda: primary producers of coral reefs. Adv. Mar. Biol.
17: 1–327.

Hoegh-Guldberg, O. et al. 2007. Coral reefs under rapid
climate change and ocean acidification. Science 318:
1737–1742.

Hinga, K.R. 2002. Effects of pH on coastal marine phy-
toplankton. Mar.Ecol.Prog.Ser. 238: 281–300.

Hossain, M.M.M. & S. Ohde. 2006. Calcification of cul-
tured Porites and Fungia under different aragonite
saturation states of seawater. Proc. 10th Int.Coral
Reef Sym. Jpn. Coral Reef Soc. 597–606.

Houghton, J.T. et al. 2001. Climate Change 2001: The Scientific
Basis. Cambridge University Press. Cambridge, UK.

Huesemann, M.H., A.D. Skillman, & E.A. Crecelius.
2002. The inhibition of marine nitrification by ocean
disposal of carbon dioxide. Mar. Pollut. Bull. 44: 142–
148.

Husebo, A. et al. 2002. Distribution and abundance of fish
in deep-sea coral habitats. Hydrobiologia 471: 91–99.

Hutchins, D.A. et al. 2007. CO2 control of Trichodesmium
N2 fixation, photosynthesis, growth rates, and el-
emental ratios: implications for past, present, and
future ocean biogeochemistry. Limnol. Oceanogr. 52:
1293–1304.

Invers, O. et al. 2001. Inorganic carbon sources for sea-
grass photosynthesis: an experimental evaluation of
bicarbonate use in species inhabiting temperate wa-
ters. J. Exp. Mar. Biol. Ecol. 265: 203–217.

Ishimatsu, A. et al. 2005. Physiological effects on fishes in
a high-CO2 world. J. Geophys. Res. 110: np.

Ishimatsu, A. et al. 2004. Effects of CO2 on marine fish:
larvae and adults. J. Oceanogr. 60: 731–741.

Ishimatsu, A. & J. Kita. 1999. Effects of environmental
hypercapnia on fish, Jpn. J. Ichthyol. 46: 1–13.

Kikkawa, T., A. Ishimatsu & J. Kita. 2003. Acute CO2
tolerance during the early developmental stages of
four marine teleosts. Environ. Toxicol. 18: 375–382.

Kikkawa, T., J. Kita, & A. Ishimatsu. 2004. Comparison
of the lethal effect of CO2 and acidification on red
sea bream (Pagrus major) during the early develop-
mental stages. Mar. Pollut. Bull. 48: 108–110.

Kleypas, J.A. et al. 1999a. Geochemical consequences of
increased atmospheric carbon dioxide on coral reefs.
Science 284: 118–120.

Kleypas, J.A. et al. 1999b. Environmental limits to coral
reef development: where do we draw the line? Am
Zool 39: 146–159.

Kleypas, J.A. et al. 2006. Impacts of Ocean Acidifica-
tion on Coral Reefs and Other Marine Calcifiers:
A Guide for Future Research, report of a workshop
held 18–20 April 2005, St. Petersburg, FL, sponsored
by NSF, NOAA, and the U.S. Geological Survey, 88
pp.

Kleypas, J.A. & C. Langdon. 2002. Overview of CO2-
induced changes in seawater chemistry. World Coral
Reefs in the New Millennium: Bridging Research
and Management for Sustainable Development. In
Proceedings of the 9th International Coral Reef
Symposium, 2. M.K. Moosa, S. Soemodihardjo, A.
Soegiarto, et al., Eds.: 1085–1089. Ministry of Envi-
ronment, Indonesian Institute of Sciences, Interna-
tional Society for Reef Studies, Bali, Indonesia.

340 Annals of the New York Academy of Sciences

Koenig, C. et al. 2000. Protection of fish spawning habitat
for the conservation of warm- temperate reef-fish
fisheries of shelf-edge reefs of Florida. Bull. Mar. Sci.
66: 593–616.

Kuffner, I.B. et al. 2008. Decreased abundance of crus-
tose coralline algae due to ocean acidification. Nature
Geoscience In Press.

Kurihara, H., S. Kato, & A. Ishimatsu. 2007. Effects of
increased seawater pCO2 on early development of
the oyster Crassostrea gigas. Aquat. Biol. 1: 91–98.

Kurihara, H., S. Shimode, & Y. Shirayama. 2004. Sub-
lethal effects of elevated concentration of CO2 on
planktonic copepods and sea urchins. J. Oceanogr. 60:
743–750.

Kurihara, H. & Y. Shirayama. 2004. Effects of increased
atmospheric CO2 on sea urchin early development.
Mar. Ecol. Prog. Ser. 274: 161–169.

Kuwatani, Y. & T. Nishii. 1969. Effects of decreased pH
of culture water on the growth of the Japanese pearl
oyster. Bull. Jap. Soc. Sci. Fish. 35: 342–350.

Langdon, C. et al. 2000. Effect of calcium carbonate sat-
uration state on the calcification rate of an experi-
mental coral reef. Global Biogeochem. Cy. 14: 639–654.

Langdon, C. et al. 2003. Effect of elevated CO2 on
the community metabolism of an experimental
coral reef. Global Biogeochem. Cy. 17: 1011, doi:
10.1029/2002GB001941.

Langdon, C. & M.J. Atkinson. 2005. Effect of elevated
pCO2 on photosynthesis and calcification of corals
and interactions with seasonal change in tempera-
ture/irradiance and nutrient enrichment. J. Geophys.
Res. 110: np.

Langer, M.R. et al. 2006. Species-specific responses of cal-
cifying algae to changing seawater carbonate chem-
istry. Geochem. Geophys. Geosyst. 7: np.

Leclercq, N., J.-P. Gattuso, & J. Jaubert. 2000. CO2 par-
tial pressure controls the calcification rate of a coral
community. Global Change Biol. 6: 329–334.

Leclercq, N. et al. 2002. Primary production, respiration,
and calcification of a coral reef mesocosm under
increased CO2 partial pressure. Limnol. Oceanogr. 47:
558–564.

Littler, M.M. & D.S. Littler. 1984. Models of tropical reef
biogenesis: the contribution of algae. Prog. Phycol. Res.
3: 323–364.

Lueker, T.J., A.G. Dickson, & C.D. Keeling. 2000. Ocean
pCO2 calculated from dissolved inorganic carbon,
alkalinity, and equations for K1 and K2: validation
based on laboratory measurements of CO2 in gas
and seawater at equilibrium. Mar. Chem. 70: 105–
119.

Marshall, A.T. & P.L. Clode. 2002. Effect of increased
calcium concentration in sea water on calcification
and photosynthesis in the scleractinian coral Galaxea
fascicularis. J. Exp. Biol. 205: 2107–2113.

Martin, C.L. & P.D. Tortell. 2006. Bicarbonate trans-
port and extracellular carbonic anhydrase activity in
Bering Sea phytoplankton assemblages: Results from
isotope disequilibrium experiments. Limnol. Oceanogr.
51: 2111–2121.

Marubini, F., C. Ferrier-Pagés & J.P. Cuif. 2003. Sup-
pression of skeletal growth in scleractinian corals by
decreasing ambient carbonate-ion concentration: a
crossfamily comparison. Proc. Roy. Soc. Lond. B 270:
179–184.

Marubini, F., H. Barnett, C. Langdon & M.J. Atkinson.
2001. Dependence of calcification on light and car-
bonate ion concentration for the hermatypic coral
Porites compressa. Mar. Ecol.-Prog. Ser. 220: 153–162.

Marubini, F. & B. Thake. 1999. Bicarbonate addition
promotes coral growth. Limnol. Oceanogr. 44: 716–
720.

McKim, J.M. 1977. Evaluation of tests with early life
stages of fish for predicting long- term toxicity. J.
Fish. Res. Board Can. 34: 1148–1154.

Medina, M.A. et al. 2006. Naked corals: skeleton loss in
Scleractinia. Proc. Natl. Acad. Sci. USA 103: 9096–
9100.

Meehl, G.A. et al. 2007. Global Climate Projections. In
Climate Change 2007: The Physical Science Basis. Contribu-

tion of Working Group I to the Fourth Assessment Report of the

Intergovernmental Panel on Climate Change. S. Solomon,
D. Qin, M. Manning, et al., Eds.: Cambridge Uni-
versity Press. Cambridge, UK, and New York, NY.

Mehrbach, C. et al. 1973. Measurement of apparent
dissociation constants of carbonic acid in seawater
at atmospheric pressure. Limnol. Oceanogr. 18: 897–
907.

Michaelidis, B.C. et al. 2005. Effects of long-term moder-
ate hypercapnia on acid-base balance and growth
rate in marine mussels Mytilus galloprovincialis,
Mar. Ecol. Prog. Ser. 293: 109–118.

Milliman, J.D. & A.W. Droxler. 1996. Neritic and pelagic
carbonate sedimentation in the marine environment:
Ignorance is not bliss. Geol. Rundsch. 85: 496–504.

Mortensen, P.B. et al. 2001. Distribution, abundance and
size of Lophelia pertusa coral reefs in mid-Norway in
relation to seabed characteristics. J. Mar. Biol. Assoc.
UK. 81: 581–597.

Mortensen, P.B. 2000. Lophelia pertusa (Scleractinia) in
Norwegian waters; distribution, growth, and asso-
ciated fauna. Ph.D. thesis, University of Bergen,
Bergen, Norway.

Mucci, A. 1983. The solubility of calcite and aragonite
in seawater at various salinities, temperatures, and
one atmosphere total pressure. Am. J. Sci. 283: 780–
799.

Mumby, P.J. et al. 2004. Mangroves enhance the biomass
of coral reef fish communities in the Caribbean.
Nature 427: 533–536.

Guinotte & Fabry: Ocean Acidification and Marine Ecosystems 341

Ohde, S. & M.M.M. Hossain. 2004. Effect of CaCO3
(aragonite) saturation state of seawater on calcifica-
tion of Porites coral. Geochem. J. 38: 613–621.

Ohde, S. & R. Van Woesik. 1999. Carbon dioxide flux
and metabolic processes of a coral reef. Bull. Mar. Sci.
65: 559–576.

Orr, J.C. et al. 2005. Anthropogenic ocean acidification
over the twenty-first century and its impact on calci-
fying organisms. Nature 437: 681–686.

Palacios, S. & R.C. Zimmerman. 2007. Response of eel-
grass Zostera marina to CO2 enrichment: possible
impacts of climate change and potential for reme-
diation of coastal habitats. Mar. Ecol. Prog. Ser. 344:
1–13.

Pane, L. et al. 2004. Summer coastal zooplankton biomass
and copepod community structure near the Italian
Terra Nova Base (Terra Nova Bay, Ross Sea, Antarc-
tica). J. Plank. Res. 26: 1479–1488.

Parrish, J.D. 1989. Fish communities of interacting
shallow-water habitats in tropical oceanic regions.
Mar. Ecol. Prog. Ser. 58: 143–160.

Petit, J.R. et al. 1999. Climate and atmospheric history
of the past 420000 years from the Vostok ice core,
Antarctica. Nature 399: 429–436.

Pollard, D.A. 1984. A review of ecological studies on sea-
grass fish communities, with particularly reference to
recent studies in Australia. Aquat. Bot. 18: 3–42.

Portner, H.O., M. Langenbuch & A. Reipschlager. 2004.
Biological impacts of elevated ocean CO2 concen-
trations: lessons from animal physiology and earth
history. J. Oceanogr. 60: 705–718.

Purcell, J.E. 2005. Climate effects on formation of jellyfish
and ctenophore blooms: a review. J. Mar Biol. Ass. UK
85: 461–476.

Purcell, J.E. & M.N. Arai. 2001. Interactions of pelagic
cnidarians and ctenophores with fishes: a review.
Hydrobiologia 451: 27–44.

Purcell, J.E., S. Uye & W.-T. Lo. 2007. Anthropogenic
causes of jellyfish blooms and their direct conse-
quences for humans: a review. Mar. Ecol. Prog. Ser.
350: 153–174.

Reed, J.K. 2002. Deep-water Oculina coral reefs of
Florida: biology, impacts and management. Hydro-
biologia 471: 43–55.

Renegar, D.A. & B.M. Riegl. 2005. Effect of nutrient
enrichment and elevated CO2 partial pressure on
growth rate of Atlantic scleractinian coral Acropora
cervicornis. Mar. Ecol.-Prog. Ser. 293: 69–76.

Reynaud, S. et al. 2003. Interacting effects of CO2 partial
pressure and temperature on photosynthesis and cal-
cification in a scleractinian coral. Global Change Biol.
9: 1660–1668.

Riebesell, U. et al. 2000. Reduced calcification of marine
plankton in response to increased atmospheric CO2.
Nature 407: 364–367.

Riebesell, U. et al. 2007. Enhanced biological carbon con-
sumption in a high CO2 ocean. Nature 450: 545–548.

Roberts, J.M., J.D. Gage & the ACES party. 2003. As-
sessing biodiversity associated with cold-water coral
reefs: Pleasures and pitfalls. Erlanger Geol. Abh. 4: 73.

Roberts, J.M., A.J. Wheeler, & A. Freiwald. 2006. Reefs
of the deep: the biology and geology of cold-water
coral ecosystems. Science 312: 543–547.

Rogers, A.D. 1999. The biology of Lophelia pertusa (Lin-
naeus 1758) and other deep- water reef-forming
corals and impacts from human activities. Int. Rev.
Hydrobiol. 84: 315–406.

Roos, A. & W.F. Boron. 1981. Intracellular pH. Physiol.
Rev. 61: 296–434.

Rost, B., U. Riebesell & S. Burkhardt. 2003. Carbon ac-
quisition of bloom-forming marine phytoplankton.
Limnol. Oceanogr. 48: 55–67.

Rost, B. & U. Riebesell. 2004. Coccolithophores and
the biological pump: responses to environmental
changes. In Coccolithophores—From Molecular Processes
to Global Impact. H.R. Thierstein & J.R. Young, Eds.:
76–99. Springer. Berlin Heidelberg.

Royal Society. 2005. Ocean Acidification Due to Increasing At-
mospheric Carbon Dioxide. Policy Document 12/05. The
Royal Society. London, UK.

Sabine, C.L. et al. 2004. The oceanic sink for anthro-
pogenic CO2. Science 305: 367–371.

Schippers, P., M. Lurling & M. Scheffer. 2004. Increase of
atmospheric CO2 promotes phytoplankton produc-
tivity. Ecol. Lett. 7: 446–451.

Schneider, K. & J. Erez. 2006. The effect of carbon-
ate chemistry on calcification and photosynthesis in
the hermatypic coral Acropora eurystoma. Limnol.
Oceanogr. 51: 1284–1293.

Sciandra, A. et al. 2003. Response of coccolithophorid
Emiliania huxleyi to elevated partial pressure of CO2
under nitrogen limitation. Mar. Ecol.-Prog. Ser. 261:
111–122.

Sheridan, P. & C. Hays. 2003. Are mangroves nursery
habitat for transient fishes and decapods? Wetlands
23: 449–458.

Siegenthaler, U. et al. 2005. Stable carbon cycle-climate
relationship during the late Pleistocene. Science 310:
1313–1317.

Silverman, J., B. Lazar & J. Erez. 2007. Effect of aragonite
saturation, temperature, and nutrients on the com-
munity calcification rate of a coral reef. J. Geophys.
Res. 112: C05004, doi:10.1029/2006JC003770.

Skirrow, G. & M. Whitfield. 1975. The effect of in-
creases in the atmospheric carbon dioxide content
on the carbonate ion concentration of surface water
at 25◦C. Limnol. Oceanogr. 20: 103–108.

Spero, H.J. et al. 1997. Effect of seawater carbonate con-
centration on foraminiferal carbon and oxygen iso-
topes. Nature 390: 497–500.

342 Annals of the New York Academy of Sciences

Stanley, G.D., Jr. 2006. Photosymbiosis and the evolution
of modern coral reefs. Science 312: 857–858.

Stanley, G.D., Jr. 2007. Ocean acidification and sclerac-
tinian corals. Science 317: 1032–1033.

Stanley, G.D., Jr. & D.G. Fautin. 2001. The Origin of
Modern Corals. Science 291: 1913–1914.

Stone, R.P. 2006. Coral habitat in the Aleutian Islands
of Alaska: depth distribution, fine-scale species as-
sociations, and fisheries interactions. Coral Reefs 25:
229–238.

Tortell, P.D. & F.M.M. Morel. 2002. Sources of inorganic
carbon for phytoplankton in the eastern Subtropical
and Equatorial Pacific Ocean. Limnol. Oceanogr. 47:
1012–1022.

Tortell, P.D., J.R. Reinfelder, & F.M.M. Morel. 1997. Ac-
tive uptake of bicarbonate by diatoms. Nature 390:
243–244.

Turley, C.M., J.M. Roberts, & J.M. Guinotte. 2007.
Corals in deep-water: will the unseen hand of ocean
acidification destroy cold-water ecosystems? Coral
Reefs 26: 445–448.

Veron, J.E. 2008. A Reef in Time: The Great Barrier Reef
from Beginning to End. Harvard University Press. Cam-
bridge, MA.

Whitfield, M. 1975. Future impact of fossil CO2 on sea–
Reply. Nature 254: 274–275.

Yates, K.K. & R.B. Halley. 2006. CO3 concentration and
pCO2 thresholds for calcification and dissolution on
the Molokai reef flat, Hawaii. Biogeosciences 3: 357–
369.

Zachos, J.C. et al. 2005. Rapid acidification of the ocean
during the Paleocene-Eocene Thermal Maximum.
Science 308: 1611–1615.

Zeebe, R.E. & D. Wold-Gladrow. 2001. CO2 in Seawater:
Equilibrium, Kinetics, Isotopes. Elsevier Oceanography
Series 65. Amsterdam. pp. 346.

Zimmerman, R.C. et al. 1997. Impacts of CO2-
enrichment on productivity and light requirements
of eelgrass. Plant Physiol. 115: 599–607.

Zondervan, I. 2007. The effects of light, macronutri-
ents, trace metals and CO2 on the production of
calcium carbonate and organic carbon in coccol-
ithophores – A review. Deep Sea Res. II 54: 521–
537.

Zondervan, I. et al. 2001. Decreasing marine biogenic
calcification: A negative feedback on rising at-
mospheric pCO2. Global Biogeochem. Cy. 15: 507–
516.

MINIREVIE

W

EFFECTS OF CLIMATE CHANGE ON GLOBAL SEAWEED COMMUNITIES1

Christopher D. G. Harley,2 Kathryn M. Anderson, Kyle W. Demes, Jennifer P. Jorve, Rebecca L. Kordas,
Theraesa A. Coyle

Department of Zoology and Biodiversity Research Centre, University of British Columbia, 6270 University Blvd, Vancouver,

British Columbia, V6T1Z4, Canada

and Michael H. Graham

Moss Landing Marine Laboratories, 8272 Moss Landing Road, Moss Landing, California, 95039, USA

Seaweeds are ecologically important primary
producers, competitors, and ecosystem engineers that
play a central role in coastal habitats ranging from kelp
forests to coral reefs. Although seaweeds are known to
be vulnerable to physical and chemical changes in the
marine environment, the impacts of ongoing and
future anthropogenic climate change in seaweed-
dominated ecosystems remain poorly understood. In
this review, we describe the ways in which changes in
the environment directly affect seaweeds in terms of
their physiology, growth, reproduction, and survival.
We consider the extent to which seaweed species may
be able to respond to these changes via adaptation or
migration. We also examine the extensive reshuffling
of communities that is occurring as the ecological
balance between competing species changes, and as
top-down control by herbivores becomes stronger or
weaker. Finally, we delve into some of the ecosystem-
level responses to these changes, including changes in
primary productivity, diversity, and resilience.
Although there are several key areas in which
ecological insight is lacking, we suggest that reasonable
climate-related hypotheses can be developed and
tested based on current information. By strategically
prioritizing research in the areas of complex
environmental variation, multiple stressor effects,
evolutionary adaptation, and population, community,
and ecosystem-level responses, we can rapidly build
upon our current understanding of seaweed biology
and climate change ecology to more effectively
conserve and manage coastal ecosystems.

Key index words: adaptation; carbon dioxide; climate
change; community structure; competition; ecophysi-
ology; ecosystem function; herbivory; marine macro-
algae; ocean acidification

Changes in global temperature and ocean chemistry
associated with increasing greenhouse gas concen-
trations are forcing widespread shifts in biological
systems. In response to warming, species ranges are
shifting toward the poles, up mountainsides, and to
deeper ocean depths (Parmesan and Yohe 2003,
Perry et al. 2005). Factors including warming and
ocean acidification are causing the reorganization
of local communities as species are added or
deleted and as interactions among species change
in importance (Wootton et al. 2008, Harley 2011).
Because greenhouse gas emission rates continue to
accelerate, the climatically forced ecological changes
that have been documented over the past half cen-
tury will likely pale in comparison to changes in the
coming decades.
Global change is, by definition, a global phenom-

enon, yet some biological systems have received
more attention than others. Although a great deal
of research has focused on systems like coral reefs
and terrestrial forests (e.g., Hoegh-Guldberg et al.
2007, Aitken et al. 2008), considerably less attention
has been devoted to seaweed-dominated ecosystems
(Wernberg et al. 2012). Like corals and trees, sea-
weeds are key habitat structuring agents that harbor
incredible biodiversity (Graham 2004, Christie et al.
2009). Seaweeds form the base of productive food
webs that include economically valuable species
(Graham 2004, Norderhaug et al. 2005) and extend
well beyond the shallow waters in which seaweeds
dwell (Harrold et al. 1998). Seaweeds are intimately
linked to human cultural and economic systems via
the provision of ecosystem goods and services rang-
ing from food to medicine to storm protection
(Rönnbäck et al. 2007).
Here, we describe how global climate change influ-

ences marine macroalgae and their associated ecosys-
tems. We begin with the physical and chemical
changes that are currently at work in the oceans, and
how these changes may impact seaweed performance
via changes in stress and resource availability. These
direct linkages from environment to organism will

1Received 24 April 2012. Accepted 17 July 2012.
2Author for correspondence: e-mail harley@zoology.ubc.ca.

J. Phycol. 48,

1064

–1078 (2012)
© 2012 Phycological Society of America
DOI: 10.1111/j.1529-8817.2012.01224.x

1064

drive species-level responses, including adaptation,
migration, and extinction. We then consider how
direct effects of climate change may modify inter-
specific interactions such as competition, herbivory,
and disease, and broaden our focus to examine
changes in whole ecosystem structure and function.
Finally, we highlight key areas where our understand-
ing is incomplete, and suggest productive avenues for
future research.

ABIOTIC CHANGE IN COASTAL MARINE ENVIRONMENTS

Rising carbon dioxide concentrations in the atmo-
sphere and in the oceans are driving a number of
important physical and chemical changes. These
include directional, global-scale trends like ocean
acidification (the shift in ocean chemistry that
includes reductions in pH and carbonate ion avail-
ability), warming, and sea-level rise, along with
regionally specific increases or decreases in wave
heights, upwelling, terrigenous nutrient runoff, and
coastal salinity. As these abiotic trends are thor-
oughly reviewed elsewhere (e.g., Feely et al. 2004,
IPCC 2007, Rabalais et al. 2009, Wang et al. 2010,
Zacharioudaki et al. 2011), we pause here only to
highlight two salient features of this suite of anthro-
pogenically forced environmental change. First, the
magnitude of change is remarkable. We have
already exceeded the maximum CO2 concentration
experienced in the last 740,000 years (Augustin
et al. 2004), and will soon exceed the range of CO2
concentrations experienced in tens of millions of
years (Pearson and Palmer 2000, IPCC 2007).
Second, the rate of abiotic change is virtually
unprecedented. The rise in CO2 concentrations and
global temperature since the industrial revolution
are 100–1000 times faster than at any point in the
past 420,000 years and are still accelerating (Hoegh-
Guldberg et al. 2007). Corresponding rates of geo-
chemical change in the oceans currently exceed
anything recorded in the last 300 million years
(Hönisch et al. 2012). Both the magnitude and the
rate of environmental change pose serious
challenges to marine species that must either
tolerate or adapt to a new ocean.

INDIVIDUAL-LEVEL RESPONSES: GAPS IN THE
ECOPHYSIOLOGICAL FRAMEWORK

Seaweed survival, growth, and reproduction are
known to vary with numerous climatically sensitive
environmental variables including temperature
(e.g., Lüning and Neushul 1978), desiccation (Davison
and Pearson 1996, Chu et al. 2012), salinity
(e.g., Steen 2004a), wave heights (Seymour et al.
1989, Graham et al. 1997), nutrient supply via
upwelling and run-off (Lobban and Harrison 1997),
pH (Kuffner et al. 2008, Martin and Gattuso 2009,
Diaz-Pulido et al. 2012), and carbon dioxide
concentration itself (Kroeker et al. 2010). To date,

our understanding of the relationship between
environmental change and the performance of indi-
vidual seaweeds is based on a loose combination of
mechanistic, physiological research, and phenome-
nological studies that correlate performance with
environmental conditions. The seaweed physiologi-
cal literature is extensive, but much of it predates
our current understanding of future environmental
scenarios, and it is not well linked to more ecologi-
cally oriented studies. Rather, climate change ecolo-
gists often make use of phenomenological studies to
make broad-brush predictions for future change.
Whether due to the lack of information or the lack
of communication across disciplines, weak or miss-
ing mechanistic linkages between predicted future
conditions and seaweed growth, reproduction, and
survival are problematic. For example, climate
change will result in novel patterns and combina-
tions of stress, and a priori predictions regarding
responses to simultaneous changes in the means
and variance of multiple environmental stressors are
difficult to make in the absence of a mechanistic
understanding of sublethal and lethal stresses in sea-
weeds. (In this review, we use the term “stress” to
denote disruptive stress sensu Davison and Pearson
(1996); stressful conditions are those that adversely
affect growth via damage and/or resource realloca-
tion associated with damage prevention and repair).
Below, we consider what is and what is not known
about the two most broadly important aspects of
environmental change, warming and ocean acidifi-
cation, with a further emphasis on variable impacts
across different algal life history stages. We then
detail some of the ways in which an incomplete
ecophysiological understanding impairs our ability to
predict seaweed responses to complex environmental
variation and multiple stressors.
Thermal ecophysiology. Temperature determines

the performance of seaweeds, and indeed all organ-
isms, at the fundamental levels of enzymatic
processes and metabolic function (reviewed in
Raven and Geider 1988, Lobban and Harrison
1997). Seaweeds have evolved biochemical and phys-
iological adaptations, including variation in the
identity and concentration of proteins and the prop-
erties of cell membranes, that enable them to
optimize their performance with respect to the tem-
peratures they encounter (Eggert 2012). Although
seaweeds are generally well adapted to their thermal
environment, they nevertheless experience tempera-
tures in nature – particularly during periods of envi-
ronmental change – that are sufficiently high or low
to result in disruptive stress in the form of cellular
and subcellular damage (reviewed in Davison and
Pearson 1996, Eggert et al. 2012). Such damage
and any reallocation of resources for protection and
repair can slow growth, delay development, and lead
to mortality (Davison and Pearson 1996). In
response, seaweeds can produce heat shock proteins
that repair or remove damaged proteins (e.g., Vayda

CLIMATE CHANGE AND SEAWEEDS 1065

M
IN

IR
E
V
IE

W

and Yuan 1994, Lewis et al. 2001). However, protein
thermal physiology is not well understood in macro-
algae (Eggert et al. 2012) and the upregulation of
heat shock protein production is only one of many
transcriptional changes that occur in seaweeds dur-
ing periods of thermal stress (Collén et al. 2007,
Kim et al. 2011). Relevant genomic, transcriptomic,
and proteomic studies are only just beginning to
scratch the surface and most links from gene
expression to organismal performance are far from
well established.

As a result of nonstressful conditions at inter-
mediate temperatures and stress at the extremes,
the relationship between temperature and most sub-
cellular, tissue-level, or whole-organism processes is
described by a hump-shaped thermal performance
curve. From colder to warmer, these curves gener-
ally rise exponentially as rates of biochemical reac-
tions increase, peak at some optimum temperature,
and then fall rapidly as the biological components
of the system become less efficient or damaged
(Kordas et al. 2011, Eggert et al. 2012). When prop-
erly parameterized across the full-temperature toler-
ance range of a species, thermal performance
curves have the potential to predict the physiologi-
cal effects of any given warming or cooling scenario
(barring any further acclimatization, adaptation, or
context-dependent surprises; see below). The effect
of a small increase in thallus temperature will be
beneficial when the initial temperature is cooler
than optimal and detrimental when it is warmer
than optimal, and the precise change in perfor-
mance can be predicted from the starting and
ending temperature values along the curve. Unfor-
tunately, the shapes of thermal performance curves
and the positions of their optima are poorly
described in most seaweeds. Although many physio-
logical and ecological studies have linked seaweed
performance to temperature, a substantial fraction
of these studies do not investigate enough tempera-
tures across a wide enough range to characterize
the underlying, nonlinear relationship between the
two. Furthermore, various physiological parameters
within an organism differ in the shape and opti-
mum temperature of their thermal performance
curves, which limits our ability to use an easily mea-
sured parameter (e.g., photosynthesis) as a proxy
for parameters that may be more ecologically
relevant (e.g., growth and reproduction). Indeed,
growth rates do tend to peak at lower temperatures
than photosynthetic rates (Eggert et al. 2012), pre-
sumably because metabolic rates increase faster than
photosynthetic rates at higher temperatures. Much
remains to be learned regarding the thermal depen-
dence of the key physiological processes that control
growth, reproduction, and survival across the full
range of temperatures experienced by an individual
in its lifetime.
Ecophysiology of ocean acidification. Carbon dioxide

concentrations in seaweed habitats are increasing

with anthropogenic emissions and, in some regions,
with intensified upwelling of CO2-enriched water
(Feely et al. 2008). As with terrestrial plants (Long
et al. 2004), it is tempting to predict that seaweeds
will benefit from the increase in inorganic carbon
concentration (Beardall et al. 1998). However, the
situation in the sea is not so simple. CO2-driven
effects on photosynthesis and growth depend on
the degree to which carbon is limiting, which in
turn varies among habitat types and among taxa.
Because CO2 diffusion rates are much higher in air
than in water, seaweeds that are exposed at low tide
and those with floating canopies at the sea-air inter-
face have greater access to CO2 (Beardall et al.
1998). However, aerial exposure does not necessar-
ily reduce the probability of carbon limitation, as
exposure at low tide can dramatically reduce rates
of carbon acquisition (Williams and Dethier 2005)
and even emersed seaweeds can benefit from
increasing atmospheric CO2 concentrations (Gao
et al. 1999, Zou and Gao 2002). Moreover, a strict
focus on CO2 in air or dissolved in water may be
misleading as not all species require environmental
CO2 as a carbon source. Most green and brown
algae (and many red algae) can also utilize bicar-
bonate (HCO3

�) by converting it to CO2 intracellu-
larly via CO2 concentrating mechanisms (CCMs; see
Raven et al. 2012 for review). Just as terrestrial C3
plants are more likely to be CO2-limited and there-
fore more likely to benefit from elevated CO2 than
C4 plants (Long et al. 2004), seaweeds lacking
CCMs are more likely to be carbon-limited and thus
more likely to benefit from additional CO2(aq). For
example, experimental addition of CO2 greatly
increased the growth rate of Lomentaria articulata,
which cannot use bicarbonate (Kübler et al. 1999),
but did not enhance photosynthetic rates of species
with CCMs or of nonbicarbonate using species that
were not carbon-limited (Cornwall et al. 2012).
However, species with CCMs did shift away from
bicarbonate and toward CO2(aq) when CO2 concen-
trations were high, which may benefit the seaweeds
by reducing the energetic costs of using CCMs
(Cornwall et al. 2012). Thus, although there may be
variation among taxa based on carbon utilization
strategy, noncalcifying seaweeds as a group will
likely respond positively to increasing global CO2
concentrations in general (see Kroeker et al. 2010).
In addition to providing carbon for photosynthesis,

anthropogenic CO2 emissions reduce seawater pH
and the saturation state of calcium carbonate. As
this increases the cost of calcification and the likeli-
hood of dissolution, calcifying organisms are partic-
ularly sensitive to elevated CO2 in seawater. Ocean
acidification is consistently related to reduced
growth rates in calcified macroalgae (Kroeker et al.
2010) and reductions in calcification rate at elevated
pCO2 have been demonstrated for crustose and
articulated coralline red algae as well as calcified
green Halimeda (Gao et al. 1993, Büdenbender

1066 CHRISTOPHER D. G. HARLEY ET AL.

M
IN
IR
E
V
IE
W

et al. 2011, Price et al. 2011). However, reduced cal-
cification at higher pCO2 did not emerge as a gen-
eral pattern in a meta-analysis of multiple seaweed
studies (Kroeker et al. 2010). This may be because
the process of calcification, and likewise the effects
of ocean acidification on calcification, varies among
seaweeds (Price et al. 2011), and many species are
able to create microclimates of chemistry favorable
for calcification regardless of ambient conditions
(Roleda et al. 2012a). It has therefore been sug-
gested that the effects of ocean acidification on cal-
cified species may be manifested as increased
dissolution rather than reduced production of cal-
cium carbonate (Roleda et al. 2012a). Reduced pH
may have important consequences for noncalcifying
taxa as well (Roleda et al. 2012b), although the
cumulative effects of climatically realistic, CO2-dri-
ven pH change on noncalcifying seaweeds remain
poorly understood.
Stress and the completion of algal life cycles. Predict-

ing true individual-level responses to climate change
in seaweeds is challenging owing to the numerous
life history stages and transitions upon which envi-
ronmental change can act (Schiel and Foster 2006).
Careful consideration of this complexity is impor-
tant because thermal optima and tolerance limits
can vary among life history stages within a species
(e.g., Fain and Murray 1982), and climate effects at
one life history stage may be magnified or offset by
impacts (or lack thereof) at other life history stages
(e.g., Ladah and Zertuche-González 2007).

To exemplify the degree to which we are ignorant
of how climate change will impact seaweeds across all
life history stages, we summarize what is known
regarding the effects of warming and elevated CO2
on one particularly well-studied species, the giant
kelp, Macrocystis pyrifera (Fig. 1). Increased tempera-
ture is generally thought to have negative effects on

spore production (Buschmann et al. 2004), germina-
tion (Buschmann et al. 2004), recruitment (Deysher
and Dean 1986a, Buschmann et al. 2004), and sporo-
phyte growth (Rothäusler et al. 2009, 2011) and con-
text-specific effects on gametogenesis depending on
the source population and degree of warming (Lüning
and Neushul 1978, Deysher and Dean 1986b,
Muñoz et al. 2004). Warming has also been linked
to mortality of spores, gametophytes, eggs, and
sporophytes (Ladah and Zertuche-González 2007).
Much less is known about the effects of increasing
CO2 concentrations. On the basis of current knowl-
edge, we can expect positive effects on gametogene-
sis and variable effects (e.g., positive effect of
increasing CO2, but negative effect of decreasing
pH) on germination (Roleda et al. 2012b). Studies
assessing the potential for interactive temperature
and CO2 effects are uncommon (see below), and
nonexistent for M. pyrifera. Thus, even for one of the
best-studied seaweeds in the world, large knowledge
gaps greatly hinder our ability to precisely predict
future changes in population growth and persistence.
The importance of variability, rates of change, and envi-

ronmental history. When predicting future ecological
patterns – and when designing experiments to test
those predictions – it is tempting to treat environ-
mental change as a steady shift in mean conditions.
However, environmental time series are complex (see
Helmuth et al. 2006 for temperature examples and
Wootton et al. 2008 for a pH example), and different
aspects of an environmental signal, including
extremes, range, and patterns of variability, will have
different biological consequences. For example,
seaweed reproduction may only occur if temperatures
drop below some threshold for a sufficiently long per-
iod of time, whereas mortality may be more closely
linked to high temperatures that exceed physiological
tolerance (Breeman 1988, Wernberg et al. 2011b).

FIG. 1. Effects of increasing temperature and CO2 on life history processes in Macrocystis pyrifera. Green boxes indicate experimental
evidence of positive effects, yellow boxes indicate negative effects, hatched boxed indicate both positive and negative (i.e., context-specific)
effects, and blank boxes represent unquantified responses owing to a lack of published information.

CLIMATE CHANGE AND SEAWEEDS 1067

M
IN
IR
E
V
IE
W

Mortality rates following short exposures to extreme
temperatures or salinities can be similar to those
found after longer exposures to less extreme condi-
tions (Forrest and Blakemore 2006). In an experi-
ment that manipulated the temporal variation of
stress, higher variability muted negative impacts of
stress on some seaweed taxa, but generated negative
impacts in others (Benedetti-Cecchi et al. 2006).
Although changes in variability can drive important
biological changes in other systems (e.g., agricul-
tural crops and forests; Southworth et al. 2000,
Giesecke et al. 2010), studies on the effects of differ-
ent magnitudes and temporal patterns of environmen-
tal variability, alone or in combination with changes in
mean conditions, are exceedingly rare for seaweeds.

Additional aspects of environmental variability
come into play when one considers that the physio-
logical limits of individuals and populations are not
constant (see below). The history of environmental
variation is often a key predictor of future success,
as an individual or population that has been
exposed to stressful conditions in the past may be
better able to cope with them in the future (Padilla-
Gamino and Carpenter 2007a). Finally, rapid envi-
ronmental changes are typically more detrimental
than slow ones, as rapid change is more likely to
outpace an organism’s ability to acclimatize or a
population’s ability to adapt (O’Connor et al.
2012). We require a better understanding of the
ecological consequences of the accelerating pace of
change in the Earth’s climate system to reduce the
probability of ecological surprises.
Multiple stressors and nonadditive effects. All of the

anthropogenically forced changes in the physical
and chemical environment are occurring simulta-
neously, and in many cases, the impact of any par-
ticular stressor on the physiology and performance
of marine macrophytes will depend upon the pres-
ence and magnitude of additional limiting or dis-
ruptive stressors. For example, the importance of –
and limiting values of – various resources are envi-
ronmentally dependent, with the degree of light
limitation at low irradiance varying with tempera-
ture (e.g., Davison et al. 1991) and the enhance-
ment of photosynthesis by elevated CO2 varying
with nutrient availability (Xu et al. 2010). The per-
cent cover of algal turfs decreased with increasing
CO2 under ambient nutrients, but the reverse was
true under elevated nutrients (Russell et al. 2009).
There are also many interactions among disruptive
stressors, including temperature, desiccation, pH,
salinity, and ultraviolet radiation. For example, in
tropical and warm-temperate crustose coralline
algae, the negative effect of warmer temperatures
on bleaching, growth rates, calcification rates, and
survival were significantly greater under conditions
of elevated CO2/reduced pH (Anthony et al. 2008,
Martin and Gattuso 2009, Diaz-Pulido et al. 2012).
The magnitude and even the direction of UV effects
depend upon temperature and CO2 (Hoffman et al.

2003, Swanson and Fox 2007, Gao and Zheng
2010). Depending on the species and life history
stage, desiccation has been shown to magnify or
reduce the effects of high temperature (e.g., Hunt
and Denny 2008, Chu et al. 2012). As of yet, it is
difficult to predict when one stressor will increase
or decrease the effect of another. There are also no
known biases toward synergistic or antagonistic
effects; in a meta-analysis of multi-stressor studies on
Fucus spp., synergistic, additive, and antagonistic out-
comes were all equally prevalent (Wahl et al. 2011).
We desperately need to incorporate more ecophysio-
logical research into a multi-stressor framework to
improve our understanding of when, where, and why
important context-dependent outcomes emerge.

POPULATION AND SPECIES-LEVEL RESPONSES:
TOLERATE, ADAPT, MOVE, OR DIE

As described above, environmental change can
elicit a wide array of responses in individual
organisms. At the species level, responses to envi-
ronmental forcing can be distilled down to a small
set of basic alternatives: (i) persistence without
acclimatization or adaptation (tolerance), (ii) per-
sistence with acclimatization or adaptation, (iii)
persistence enabled by migration to remain within
some particular climatic niche and (iv) extinction.
In this section, we devote our discussion to potential
roles of acclimatization and adaptation in facilitat-
ing local persistence, to changes in seaweed distribu-
tional patterns, and to the potential for seaweed
extirpations.
Scope for acclimatization and adaptation. There is a

rich literature on seaweed acclimation (an individ-
ual-level response to experimental manipulation of
the environment), acclimatization (an individual-
level response to natural variation in the environ-
ment), and local adaptation (a population-level
response to natural environmental variation) as a
consequence of variation in temperature, salinity,
light, and wave forces (Lüning 1990, Lobban and
Harrison 1997, Eggert et al. 2012). Appropriately
acclimated/acclimatized individuals or adapted pop-
ulations may be better able to withstand coming
environmental change. For example, warm-accli-
mated Saccharina latissima sporophytes required less
light to achieve maximum photosynthetic rates and
were more photosynthetically efficient at high tem-
peratures (Davison et al. 1991), and warm-accli-
mated Fucus vesiculosus embryos were more likely to
survive periods of thermal stress (Li and Brawley
2004). Although many species can acclimate to envi-
ronmental changes, some algal populations or
species may be less able to do so than others. For
example, some tropical species and subpopulations
appear to have limited scope for acclimation
relative to their temperate counterparts, presum-
ably due to reduced environmental variability in
tropical habitats (Padilla-Gamino and Carpenter

1068 CHRISTOPHER D. G. HARLEY ET AL.

M
IN
IR
E
V
IE
W

2007a,b). Regardless, the use of appropriately accli-
mated/acclimatized individuals is a prerequisite for
realistic climate change experiments; otherwise,
short-term measurements may not reflect true
long-term responses.

Unlike acclimation, relatively little is known about
the degree to which evolutionary adaptation may
“rescue” seaweed species in the face of environmen-
tal change. The existence of local ecotypes (e.g.,
Breeman 1988) clearly indicates that adaptation is
possible, and there is evidence to suggest that sea-
weeds can evolve and even speciate fairly rapidly (in
~400 years in the case of Fucus radicans) when suffi-
cient selective pressure is applied by the environ-
ment (Pereyra et al. 2009). However, the degree to
which most multi-cellular marine species such as
seaweeds, their competitors, and their consumers
can adapt over climate change-relevant time scales
(<100 years) is largely unknown (but see Sunday et al. 2011). Understanding the extent to which spe- cies will acclimatize or adapt to environmental change is crucial for predicting future ecological change. Distributional shifts and the threat of extinction. Because

environmental conditions directly and indirectly
influence seaweed distributional patterns at a vari-
ety of scales (Breeman 1988), changes in the envi-
ronment will result in changes in seaweed
distributions. Some of the most readily detectable
changes are at local (site) scales, where environ-
mental change can result in shifts in the vertical
distribution (often termed zonation) of intertidal
and subtidal seaweeds. Sea-level rise will result in a
general upward shift of benthic communities en
masse, although the accompanying changes in the
relative availability of appropriate substratum types
and orientations at specific shore levels (e.g., on
shores with wave cut platforms and cliffs) may drive
changes in relative algal abundance (Vaselli et al.
2008). However, zonation patterns are determined
by far more than just position relative to mean sea
level. The upper limit of intertidal seaweeds is
related to thermal and desiccation stress during
low tide (e.g., Harley 2003), and long-term
increases in air temperature have resulted in down-
shore shifts in the upper limit of some species
(Harley and Paine 2009). The depth range of sub-
tidal kelps also depends critically on environmental
factors such as temperature, water motion, and
water transparency (Graham et al. 2007), and cli-
mate-related changes in these factors are predicted
to reduce the depth range of kelp forests (Méléder
et al. 2010). When the upper and lower depth lim-
its of a species are set by different agents (e.g.,
thermal stress, light availability, consumers; Harley
2003, Graham et al. 2007, Méléder et al. 2010), cli-
mate change can result in certain species being
squeezed out of the system entirely (Harley 2011).

Climate change will drive distributional shifts at
larger, alongshore scales as well. Increased storm

frequency could restrict vulnerable species to pro-
tected shorelines, and changes in salinity may allow
seaweeds to penetrate further into, or be forced fur-
ther out of, estuarine embayments and lagoons. The
most notable large-scale changes, however, are those
occurring across latitudinal temperature gradients.
Drastic population declines and even local extinc-
tions have been documented at the warm (lower lati-
tude) end of species’ biogeographic ranges during
periods of warming (e.g., Serisawa et al. 2004).
Range retraction at low latitudes can be offset by
expansion into higher latitudes, as in western Europe
where warm-water species have expanded northward
(Lima et al. 2007). However, such expansions may
not be a sustainable escape mechanism for species
along coastlines with significant geomorphic barri-
ers, such as the end of a continent. For example,
poleward migration of seaweed species has been
observed along the east and west coasts of Australia
since 1940, but because there is no suitable habitat
within the range of most species’ dispersal abilities
further south, continued poleward retreat may
result in numerous extinctions as species ‘fall off
the map’ (Wernberg et al. 2011a). Indeed, extinc-
tions have already been documented for several
marine macroalgae, although the relative contribu-
tion of environmental change to these losses
remains poorly understood (Brodie et al. 2009).

COMMUNITY-LEVEL RESPONSES: INTERSPECIFIC
INTERACTIONS AND INDIRECT EFFECTS

Ecological change in coastal ecosystems reflects
the combined influence of direct environmental
impacts on individual species and indirect effects
mediated by changes in interspecific interactions
(Harley et al. 2006). We first describe some of the
ways that competitive, trophic, and symbiotic rela-
tionships are likely to change in seaweed systems,
and then discuss the consequences of these changes
for entire ecosystems in the following section.
Competitive relationships. Seaweeds compete for

nutrients, light, and space for attachment, and their
relative success at acquiring these resources in the
presence of other photo-autotrophs (or sessile inver-
tebrates, in the case of space) depends upon both
resource availability and environmental stress. The
availability of several resources (e.g., CO2, nitrate,
ammonium) is changing due to human activities,
and the effects of changing resource supply will
depend on the magnitude and direction of these
changes and the degree to which these resources
limit algal growth and competitive ability. Increasing
nitrogen loading tends to favor fast-growing species
with high nitrogen requirements. In some cases, this
may lead to competitive dominance by weedy taxa
(Steen 2004b, Vermeij et al. 2010) and – should
nutrients trigger a phytoplankton bloom – shading-
out of benthic seaweeds by phytoplankton (Kava-
naugh et al. 2009). In other cases, higher nitrogen

CLIMATE CHANGE AND SEAWEEDS 1069

M
IN
IR
E
V
IE
W

merely allows for the persistence of nitrogen-limited
taxa and thus enhances algal diversity (Bracken and
Nielsen 2004). Elevated CO2(aq) will differentially
affect seaweeds depending on their carbon capture
strategy. The influence of elevated CO2(aq) on sea-
weeds with carbon concentrating mechanisms, such
as kelps, is highly light-dependent, and the overall
effect of rising CO2(aq) on kelp competitive ability
remains unclear (Hepburn et al. 2011). On the
other hand, for species that rely on aqueous CO2, like
turf-forming rhodophytes in New Zealand, elevated
CO2(aq) should differentially favor their growth,
which may in turn enhance their competitive ability
(Hepburn et al. 2011).

Changes in the severity of environmental stressors
(e.g., temperature, pH, salinity, wave forces) will
also affect the outcome of competitive relationships.
In some cases, environmental extremes remove
otherwise dominant competitors, allowing subordi-
nate species to persist (Sousa 1979) and facilitating
the establishment of non-native taxa (Miller et al.
2011). Stress need not be lethal to influence the
outcome of competitive interactions (Davison and
Pearson 1996). For example, many important com-
petitors for space are calcified taxa such as crustose
coralline algae, corals, and mussels, and inhibition
of growth by reduced pH likely contributes to
increasing fleshy algal competitive dominance over
these groups (Wootton et al. 2008, Diaz-Pulido et al.
2011, Hepburn et al. 2011). The effects of rising
temperatures may increase or decrease competition,
and even change competitive interactions into facili-
tative ones. Elevated temperature increased the
competitive impacts of Enteromorpha on two species
of Fucus (Steen 2004b). Conversely, the effects of
intertidal Ascophyllum nodosum on understory barna-
cles, like the effects of subtidal Ecklonia radiata on
E. radiata recruits, shifted from negative (competi-
tive) to net positive (facilitative) at high tempera-
tures (Leonard 2000, Wernberg et al. 2010).
Herbivory. Herbivores are key structuring agents

in algal communities, influencing everything from
the survival of individual seaweeds to total algal bio-
mass and diversity (Lubchenco and Gaines 1981).
The outcomes of pairwise plant-herbivore interac-
tions depend on characteristics of both the alga and
the herbivore, including the palatability of seaweeds,
the per capita consumption rates of herbivores, and
the individual and population growth rates and
overall abundance of both. Abiotic factors associated
with climate change are known to impact all of
these attributes.

The amount of algal tissue that an herbivore can
or will consume depends on the degree of morpho-
logical or chemical defense and on aspects of
nutritional quality such as the C:N ratio (Duffy and
Hay 1990, Van Alstyne et al. 2009). Elevated temper-
ature reduced herbivore defenses in F. vesiculosus
(Weinberger et al. 2011), and changes in nutrient
availability have been shown to alter algal palatability

(e.g., Hemmi and Jormalainen 2002). Calcium car-
bonate in algal tissue is an important anti-herbivore
defense (Hay et al. 1994) and ocean acidification
may have dramatic impacts on the palatability of cal-
cified seaweeds via reduced calcification or
increased dissolution. Although elevated CO2 would
be expected to increase C:N ratios in noncalcified
taxa, the effects of elevated CO2 on seaweed palat-
ability lags far behind our understanding of such
effects in phytoplankton and terrestrial plants.
Climate change will also have direct effects on

herbivores that will cascade down to primary pro-
ducers. Several field studies suggest that warming
sea surface temperatures are associated with
increases in important herbivore populations and
concomitant declines in certain algal species (Hart
and Scheibling 1988, Ling 2008, Hernandez et al.
2010). Although warming may benefit some grazer
populations, ocean acidification is likely to be gen-
erally detrimental to many invertebrate herbivores,
particularly heavily calcified species such as sea
urchins and molluscs (Dupont et al. 2010, Crim
et al. 2011). Volcanic CO2 vent systems provide a
glimpse into this future; areas of reduced pH near
CO2 seeps are associated with reductions in urchin
and shelled gastropod abundance, and the success
of Padina spp. (despite a reduction of calcium car-
bonate in the thalli) and of highly palatable Sargas-
sum vulgar near CO2 vents has been attributed to
the absence of urchin grazers (Porzio et al. 2011,
Johnson et al. 2012). The impacts of ocean acidifi-
cation on other herbivorous taxa, notably crusta-
ceans and fish, appear to be relatively minor
(Kroeker et al. 2010).
Although useful as a starting point, changes to algal

and invertebrate performance or population size in
isolation cannot fully predict changes in the impor-
tance of herbivory. Rather, the overall impact of her-
bivory depends upon the balance of production and
consumption of algal tissue. Metabolic theory
predicts that metabolic rate and scope for activity –
which in ectothermic herbivores determine the
demand for and ability to acquire food, respectively,
– increase more quickly with temperature than algal
photosynthetic rate and thus primary production
(O’Connor 2009). As a result of these different tem-
perature–performance relationships, experimental
warming increased the relative importance of
amphipod grazing and decreased algal biomass
despite generally positive direct effects of warming
on algal growth (O’Connor 2009). Such rate-depen-
dent generalizations fall apart, however, when
abiotic conditions become stressful, and stress differ-
entially reduces the performance of one or more
of the interacting species (Kordas et al. 2011).
One trophic level or the other is often dispropor-
tionately susceptible to stress associated with
extremes in temperature, salinity, and wave forces
(Cubit 1984, Elfwing and Tedengren 2002, Taylor
and Schiel 2010), making seaweeds relatively safe or

1070 CHRISTOPHER D. G. HARLEY ET AL.

M
IN
IR
E
V
IE
W

relatively vulnerable to grazing at certain places and
times. As the environment changes, the times and
places that seaweeds are most, or least, impacted by
herbivory will change as well (see e.g., Vinueza et al.
2006).
Epibionts, endophytes, and pathogens. Seaweeds live

in constant association with a variety of microbes,
fungi, animals, and other algae that live on or in their
tissues. Of these relationships, the ecological role of
epibionts is particularly well-studied, with effects on
seaweed hosts ranging from reduced growth and
reproduction to increased risk of mechanical break-
age (e.g., Dantonio 1985). In kelp beds in eastern
Canada, outbreaks of non-native epiphytic bryozoans
are triggered by warming events, and these outbreaks
have led to drastic reductions in the percent cover of
habitat-forming Saccharina longicruris (Scheibling and
Gagnon 2009, Saunders et al. 2010). In cases such as
this, where bryozoan epibionts increase the risk of
frond breakage (Krumhansl et al. 2011), any local
increase in storminess may act synergistically with
warming and infestations by epibionts.

In contrast to epibionts, the ecology of seaweed
endophytes and diseases is poorly understood, par-
ticularly with regard to climate change (Eggert et al.
2010, Gachon et al. 2010). There is, however,
mounting evidence that warming will negatively
impact seaweeds by facilitating bacterial infections
(Campbell et al. 2011, Case et al. 2011). Departures
from optimal salinity and irradiance can also make
seaweeds more susceptible to bacterial disease, as
evidenced by experiments and observations on
aquaculture species such as Kappaphycus alvarezii
(Largo et al. 1995). Although evidence for primarily
negative pathogen-mediated effects of environmen-
tal change is slowly accumulating, the generality and
future magnitude of such negative effects remain
essentially unknown.

SHIFTS IN COMMUNITY STRUCTURE

AND ECOSYSTEM FUNCTION

Environmental change, coupled with shifts in spe-
cies interactions and the shuffling of species distri-
butions, will culminate in potentially far-reaching
changes in community structure and ecosystem
function (Harley et al. 2006). Because the responses
of benthic assemblages are often highly idiosyn-
cratic, generalizations and specific predictions are
fraught with uncertainty. While the future states of
marine ecosystems are far from certain, neither are
they completely unforeseeable. Several predictions
and testable hypotheses can be developed around
our current understanding of seaweed-dominated,
or potentially seaweed-dominated, ecosystems.

One system for which specific predictions have
been made is the rocky intertidal zone in Britain,
where the effects of climate change have been
considered in great detail and where relevant long-
term datasets exist (Hawkins et al. 2008, 2009).

Wave-protected and semi-exposed British shores are
typically dominated by large fucoid algae, which are
dominant competitors for primary space as well as
ecosystem engineers that provide cool, moist micro-
habitats for associated species (e.g., Schonbeck and
Norton 1980, Thompson et al. 1996). Rising air
temperatures and increasing wave exposure will
directly reduce fucoid canopies via lethal physiologi-
cal and hydrodynamic stress (Hawkins et al. 2009).
The appearance and/or increased abundance of
warm-water herbivores is expected to further reduce
algal cover (Mieszkowska et al. 2006, Hawkins et al.
2009). A more diverse grazer assemblage, coupled
with the replacement of a structurally complex,
cold-water barnacle species with a structurally sim-
ple, warm-water barnacle species, will reduce oppor-
tunities for fucoid size escapes from microscopic
stages and thus inhibit the regrowth of the algal
canopy (Hawkins et al. 2008). The net result is a
decline in subcanopy habitat and a reduction in
benthic primary productivity. These changes are
predicted to reduce the abundance of many inverte-
brates that rely on cool moist microhabitats and
decrease invertebrate production in the algal detri-
tus food web of the strand line; both of these effects
may negatively impact birds that forage on these
invertebrate resources (Kendall et al. 2004).
Like intertidal fucoids, subtidal kelps provide hab-

itat structure for numerous species, including many
that are economically important (Graham 2004).
Kelp forests in Australia, like fucoid communities in
Britain, are experiencing range expansions and con-
tractions of both seaweeds and important herbivores
in association with warming temperatures (Ling
2008, Wernberg et al. 2011a). In this system, there
is also evidence that ocean acidification will result
in important shifts in community structure. Experi-
mental increases of temperature and CO2 increased
the biomass of algal turfs (Connell and Russell
2010). Enhanced cover and biomass of turf-forming
algae associated with elevated CO2 occurred at the
expense of coralline crusts, although the magnitude
of this shift depended on nutrient and light levels
(Russell et al. 2009, 2011). Increasing dominance of
turfs in response to rising CO2 may in turn inhibit
kelp recruitment, which could cause or maintain
phase shifts from kelps to turfs (Connell and Russell
2010). However, kelp canopy can, to some degree,
inhibit the positive effects of elevated CO2 on turfs,
suggesting that intact kelp forests may be resistant,
but not resilient, to phase shifts to turf-dominated
communities (Falkenberg et al. 2012).
The degree to which results from southwestern

Australia will generalize to other kelp systems, such
as those under strong top-down control, is unclear.
Limited information suggests that elevated CO2 has
variable but often positive effects on kelps like Nereo-
cystis luetkeana and M. pyrifera (Thom 1996, Swanson
and Fox 2007, Roleda et al. 2012b), but negative
effects on crustose coralline algae (CCA; see above)

CLIMATE CHANGE AND SEAWEEDS 1071

M
IN
IR
E
V
IE
W

and important kelp consumers such as sea urchins
(Dupont et al. 2010, Reuter et al. 2011). Because
urchins benefit CCA by preventing overgrowth by
other seaweeds, and CCA benefit urchins by provid-
ing settlement cues, models suggest that reductions
in either taxa may result in a positive feedback loop
(Baskett and Salomon 2010). In contrast to the neg-
ative effects of ocean acidification on herbivores,
the effects of warming may be largely positive for
herbivores (Hart and Scheibling 1988, Ling 2008,
Hernandez et al. 2010). A long-term study of warm-
ing associated with power plant thermal effluent in
central California has shown that a ~3.5°C increase
in temperature results in increases in herbivore
abundance, shifts from cold-water N. luetkeana to
warmer water M. pyrifera canopies, and a replace-
ment of understory kelps by foliose red algae
(Schiel et al. 2004). A reasonable working hypothe-
sis, therefore, is that kelps in the California Current
system, particularly in the southern portions of their
range, may respond positively to the direct and indi-
rect effects of acidification, but negatively to the
direct and indirect effects of warming (Fig. 2). The
relative balance between these opposing forces,
particularly in systems with complex trophic and
competitive relationships, remains uncertain.

Tropical coral reefs are also quite sensitive to
climate change (Fig. 3). In these systems, corals and
CCA currently dominate in part because present-day
conditions are relatively conducive to calcification
and in part because herbivores benefit the slow-
growing calcifiers – despite some negative impacts
of bioerosion – by keeping fleshy macroalgae in
check (Hoegh-Guldberg et al. 2007, O’Leary and
McClanahan 2010). However, future changes in
ocean climate are predicted to destabilize coral reef
ecosystems, resulting in phase shifts from coral-dom-
inated reefs to benthic systems dominated by fleshy
macroalgae (Hoegh-Guldberg et al. 2007, Anthony
et al. 2011, Diaz-Pulido et al. 2011). Fleshy macro-
algae are positively affected by increased CO2 (Kuff-
ner et al. 2008, but see Jokiel et al. 2008), which
along with elevated nutrients may increase their
competitive ability. In contrast, reef-building corals
appear to be in serious trouble due to the influence
of climatic stressors; warming ocean waters are asso-
ciated with mass coral bleaching events, increased
pCO2 decreases coral calcification and growth and
increases dissolution, and storms cause physical
damage to weakened reef structures (Hoegh-Guld-
berg et al. 2007). Other calcified habitat-forming
reef organisms, such as CCA and the green alga
Halimeda spp., are also expected to do poorly when
pH and calcium carbonate saturation drop and tem-
peratures rise (Kuffner et al. 2008, Price et al. 2011,
Diaz-Pulido et al. 2012). Indeed, coralline algae are
relatively rare or absent from both tropical and
temperate sites with naturally occurring carbon
dioxide seeps (Hall-Spencer et al. 2008, Fabricius
et al. 2011, Porzio et al. 2011). CCA are particularly

important as they are the “cement” that helps hold
coral reefs together and provide important settle-
ment surfaces for coral larvae; the loss of these
crusts is predicted to expedite phase shifts on

FIG. 2. Future ecological scenarios for temperate kelp forests.
Solid and dashed arrows represent direct and indirect effects of
one species on another, respectively (the flow of energy via tro-
phic interactions is omitted for clarity). Faded icons represent
functional groups that may still be present but play a strongly
reduced ecological role. Relative to present-day conditions (upper
panel), future warming (middle panel) will favor grazers and
have direct and indirect negative impacts on canopy-forming
kelps. Future increases in CO2 (lower panel) will have strong neg-
ative effects on crustose coralline algae and positive effects on
noncalcified seaweeds both directly via improved growth and indi-
rectly via reduced consumption by calcified herbivores. The com-
bined impacts of simultaneous warming and acidification in a
more realistic climate change scenario remain poorly understood.
See text for details.

1072 CHRISTOPHER D. G. HARLEY ET AL.

M
IN
IR
E
V
IE
W

tropical reefs (Diaz-Pulido et al. 2007). The key to
triggering a phase shift from corals to fleshy macroal-
gae, however, may rest with the herbivores (Fig. 3).
In areas where herbivore biomass can be maintained,
the shift away from coral-dominated systems can be
delayed, although perhaps not prevented indefinitely
(Hughes et al. 2003, Hoegh-Guldberg et al. 2007,

Buddemeier et al. 2011). Should structurally com-
plex corals be replaced with fleshy macroalgae, a
considerable loss of biodiversity would result (Hoegh-
Guldberg et al. 2007).
Ecosystem shifts, such as those described above,

may occur rapidly once a system has been pushed
beyond some threshold or tipping point (Scheffer
et al. 2001). In some cases, the behavior of the sys-
tem may not change over a wide range of progres-
sive impairment (e.g., biomass removal or species
loss), only to shift suddenly once a threshold is
crossed (Speidel et al. 2001, Davies et al. 2011).
Catastrophic phase shifts, as have been observed in
kelp forests and on coral reefs, are facilitated by
losses of resilience associated with changes in
resource supply, food web structure, and distur-
bance frequency (Folke et al. 2004), all of which are
altered by CO2-induced environmental change.
Catastrophic shifts are often difficult to anticipate,
as relevant environmental thresholds may lie at or
beyond the range of historical variation, but within
the range of near-future environmental conditions.
Should environmental conditions return to below-
threshold values, recovery may proceed quickly,
slowly, or not at all (Folke et al. 2004), and the recovery
trajectory may differ considerably from the original per-
turbation trajectory (e.g., Baskett and Salomon 2010).

FUTURE DIRECTIONS: ADDRESSING THE BIG UNKNOWNS

Although a great deal of progress has been made
in recent years, there are still significant gaps in our
understanding which hamper our ability to predict
the outcomes of global change in seaweed-domi-
nated systems. Some of the most important areas in
which we lack a general or even basic understanding
include (i) the importance of rates, timing, magni-
tude, and duration of environmental change,
(ii) non-additive effects of multiple stressors,
(iii) population-level implications of variable envi-
ronmental impacts among life-history stages,
(iv) the scope for population- or species-level adap-
tation to environmental change and (v) ecological
responses at the level of communities and ecosys-
tems, including tipping points and sudden phase
shifts. With regard to uncertainties in the nature of
environmental forcing, we require additional ecophys-
iological and ecomechanical studies – especially ones
that move beyond single-factor ANOVA designs –
and further development in the emerging field of
ecological genomics to identify biological responses
to key environmental drivers or combinations of
drivers. Of particular use would be an ecophysiologi-
cal framework from which the impacts of multiple
stressors could be predicted a priori (Pörtner and
Farrell 2008). Once understood, these drivers can
be incorporated into demographic models to better
describe and predict changes in population growth
or decline. Although species-level research on sea-
weeds, at least with regard to climate change, lags

FIG. 3. Future ecological scenarios for tropical coral reefs.
Arrows and shading as in Fig. 2. Relative to the present day
(upper panel), the combination of warming and ocean acidificat-
ion will reduce the dominance of calcified taxa such as crustose
coralline algae and corals (middle and lower panels). However,
the likelihood of fleshy macroalgae rising to dominance and out-
competing the calcified taxa depends upon whether they are sup-
pressed by herbivores (as may happen in a marine protected
area, middle panel) or not (as may happen on a heavily fished
reef, lower panel). See text for details.

CLIMATE CHANGE AND SEAWEEDS 1073

M
IN
IR
E
V
IE
W

behind similar work in terrestrial environments
(e.g., Aitken et al. 2008), there is no reason that
phycologists could not model a research program
based on the successes of terrestrial botanists, foresters,
and agricultural scientists. As for community and
ecosystem-level change, researchers can make rapid
progress by focusing on ecological dominants (e.g.,
kelps) and strong interactors (e.g., sea urchins) as a
starting point. Individual pieces of the ecological
puzzle can then be interlinked with mathematical
models and ground-truthed in areas where environ-
mental conditions already approximate future pro-
jections (e.g., volcanic CO2 vents and power plant
thermal effluent plumes).

Seaweed beds, coral reefs, and other coastal eco-
systems provide trillions of dollars of ecosystem
goods and services every year (Costanza et al. 1997),
and the degradation of these systems will have far-
reaching consequences for human societies. Devel-
oping accurate predictions for the ecological effects
of climate change in seaweed-dominated systems is
therefore a high priority, as it will be invaluable for
effective conservation and management. The climate
change scenario leading from healthy coral reefs to
degraded macroalgal beds is an excellent example
of an ecological prediction that can be used to dic-
tate management priorities. Although warming and
ocean acidification are beyond our control in the
near term, we can manage for coral reef resilience
by conserving herbivore diversity and abundance
and reducing nutrient loads (Hoegh-Guldberg et al.
2007). In some parts of the Caribbean, this strategy
appears to work in practice; following high tempera-
ture and hurricane disturbances, coral recovery rates
were higher in protected areas where algal cover was
more effectively controlled by herbivores (Mumby
and Harborne 2010). There is high yet largely
untapped potential for similarly feasible local-scale
management options in a wide variety of seaweed-
dominated coastal ecosystems that are undergoing
major ecological reorganization in response to
anthropogenic change (e.g., Russell et al. 2009).
Identifying the leverage points where conservation
and management practices are most effective should
continue to be a major focus of ecological research.

We thank S. Dudgeon and two anonymous reviewers for their
constructive criticisms. M. Mach kindly provided the illustra-
tions. During the writing of this paper, CH was supported in
part by the Killam Trusts and by the Hakai Network for
Coastal People, Ecosystems, and Management at Simon
Fraser University. Funding was provided by the NSF through
grant OCE 0752523 to M. Graham and C. Harley.

Aitken, S. N., Yeaman, S., Holliday, J. A., Wang, T. L. & Curtis-
McLane, S. 2008. Adaptation, migration or extirpation: climate
change outcomes for tree populations. Evol. Appl. 1:95–111.

Anthony, K. R. N., Kline, D. I., Diaz-Pulido, G., Dove, S. & Hoegh-
Guldberg, O. 2008. Ocean acidification causes bleaching and
productivity loss in coral reef builders. Proc. Natl. Acad. Sci.
USA 105:17442–6.

Anthony, K. R. N., Maynard, J. A., Diaz-Pulido, G., Mumby, P. J.,
Marshall, P. A., Cao, L. & Hoegh-Guldberg, O. 2011. Ocean
acidification and warming will lower coral reef resilience.
Global Change Biol. 17:1798–808.

Augustin, L., Barbante, C., Barnes, P. R. F., Barnola, J. M., Bigler, M.,
Castellano, E., Cattani, O. et al. 2004. Eight glacial cycles from
an Antarctic ice core. Nature 429:623–8.

Baskett, M. L. & Salomon, A. K. 2010. Recruitment facilitation
can drive alternative states on temperate reefs. Ecology
91:1763–73.

Beardall, J., Beer, S. & Raven, J. A. 1998. Biodiversity of marine
plants in an era of climate change: some predictions based
on physiological performance. Bot. Mar. 41:113–23.

Benedetti-Cecchi, L., Bertocci, I., Vaselli, S. & Maggi, E. 2006.
Temporal variance reverses the impact of high mean
intensity of stress in climate change experiments. Ecology
87:2489–99.

Bracken, M. E. S. & Nielsen, K. J. 2004. Diversity of intertidal
macroalgae increases with nitrogen loading by invertebrates.
Ecology 85:2828–36.

Breeman, A. M. 1988. Relative importance of temperature and
other factors in determining geographic boundaries of
seaweeds: experimental and phenological evidence. Helgol.
Wiss. Meeresunters 42:199–241.

Brodie, J., Andersen, R. A., Kawachi, M. & Millar, A. J. K. 2009.
Endangered algal species and how to protect them.
Phycologia 48:423–38.

Buddemeier, R. W., Lane, D. R. & Martinich, J. A. 2011.
Modeling regional coral reef responses to global warming
and changes in ocean chemistry: Caribbean case study. Clim.
Change 109:375–97.

Büdenbender, J., Riebesell, U. & Form, A. 2011. Calcification of
the Arctic coralline red algae Lithothamnion glaciale in
response to elevated CO2. Mar. Ecol. Prog. Ser. 441:79–87.

Buschmann, A. H., VaAquez, J., Osorio, P., Reyes, E., Filun, L.,
Hernandez-Gonzalez, M. C. & Vega, A. 2004. The effect of
water movement, temperature and salinity on abundance
and reproductive patterns of Macrocystis spp. (Phaeophyta) at
different latitudes in Chile. Mar. Biol. 145:849–62.

Campbell, A. H., Harder, T., Nielsen, S., Kjelleberg, S. &
Steinberg, P. D. 2011. Climate change and disease: bleaching
of a chemically defended seaweed. Global Change Biol.
17:2958–70.

Case, R. J., Longford, S. R., Campbell, A. H., Low, A., Tujula, N.,
Steinberg, P. D. & Kjelleberg, S. 2011. Temperature
induced bacterial virulence and bleaching disease in a
chemically defended marine macroalga. Environ. Microbiol.
13:529–37.

Christie, H., Norderhaug, K. M. & Fredriksen, S. 2009.
Macrophytes as habitat for fauna. Mar. Ecol. Prog. Ser.
396:221–33.

Chu, S. H., Zhang, Q. S., Liu, S. K., Tang, Y. Z., Zhang S. B.,
Lu Z. C. & Yu Y. Q. 2012. Tolerance of Sargassum thunbergii
germlings to thermal, osmotic and desiccation stress. Aq. Bot.
96:1–6.

Collén, J., Guisle-Marsollier, I., Léger, J. J. & Boyen, C. 2007.
Response of the transcriptome of the intertidal red seaweed
Chondrus crispus to controlled and natural stresses. New
Phytol. 176:45–55.

Connell, S. D. & Russell, B. D. 2010. The direct effects of
increasing CO2 and temperature on non-calcifying
organisms: increasing the potential for phase shifts in kelp
forests. Proc. R. Soc. Lond. B Biol. Sci. 277:1409–15.

Cornwall, C. E., Hepburn, C. D., Pritchard D., Currie, K. I.,
McGraw, C. M., Hunter, K. A. & Hurd, C. L. 2012. Carbon
use strategies in macroalgae: differential responses to
lowered pH and implications for ocean acidification.
J. Phycol. 48:137–44.

Costanza, R., d’Arge, R., de Groot, R., Farber, S., Grasso, M.,
Hannon, B., Limburg, K. et al. 1997. The value of the world’s
ecosystem services and natural capital. Nature 387:253–60.

Crim, R. N., Sunday, J. M. & Harley, C. D. G. 2011. Elevated
seawater CO2 concentrations impair larval development and

1074 CHRISTOPHER D. G. HARLEY ET AL.

M
IN
IR
E
V
IE
W

reduce larval survival in endangered northern
abalone (Haliotis kamtschatkana). J. Exp. Mar. Biol. Ecol. 400:
272–7.

Cubit, J. D. 1984. Herbivory and the seasonal abundance of algae
on a high intertidal rocky shore. Ecology 65:1904–17.

Dantonio, C. 1985. Epiphytes on the rocky intertidal red alga
Rhodomela larix (Turner) C Agardh – negative effects on the
host and food for herbivores. J. Exp. Mar. Biol. Ecol. 86:
197–218.

Davies, T. W., Jenkins, S. R., Kingham, R., Kenworthy, J.,
Hawkins, S. J. & Hiddink, J. G. 2011. Dominance, biomass
and extinction resistance determine the consequences of
biodiversity loss for multiple coastal ecosystem processes.
PLoS ONE 6:e28362.

Davison, I. R., Greene, R. M. & Podolak, E. J. 1991. Temperature
acclimation of respiration and photosynthesis in the brown
alga Laminaria saccharina. Mar. Biol. 110:449–54.

Davison, I. R. & Pearson, G. A. 1996. Stress tolerance in intertidal
seaweeds. J. Phycol. 32:197–211.

Deysher, L. E. & Dean, T. A. 1986a. In situ recruitment of
sporophytes of the giant kelp, Macrocystis pyrifera (L) C.A.
Agardh – effects of physical factors. J. Exp. Mar. Biol. Ecol.
103:41–63.

Deysher, L. E. & Dean, T. A. 1986b. Interactive efects of light and
temperature on sporophyte production in the giant kelp
Macrocystis pyrifera. Mar. Biol. 93:17–20.

Diaz-Pulido, G., Anthony, K. R. N., Kline, D. I., Dove, S. &
Hoegh-Guldberg, O. 2012. Interactions between ocean
acidification and warming on the mortality and dissolution
of coralling algae. J. Phycol. 48:32–9.

Diaz-Pulido, G., Gouezo, M., Tilbrook, B., Dove, S. & Anthony, K.
R. N. 2011. High CO2 enhances the competitive strength of
seaweeds over corals. Ecol. Lett. 14:156–62.

Diaz-Pulido, G., McCook, L. J., Larkum, A. W. D., Lotze, H. K.,
Raven, J. A., Schaffelke, B., Smith, J. E. & Steneck, R. S.
2007. Vulnerability of macroalgae of the Great Barrier Reef
to climate change. In Johnson, J. E. & Marshall, P. A. [Eds.]
Climate Change and the Great Barrier Reef. Great Barrier Reef
Marine Park Authority and Australian Greenhouse Office,
Townsville, pp. 154–92.

Duffy, J. E. & Hay, M. E. 1990. Seaweed adaptations to herbivory –
Chemical, structural, and morphological defenses are often
adjusted to spatial or temporal patterns of attack. Bioscience
40:368–75.

Dupont, S., Ortega-Martinez, O. & Thorndyke, M. 2010. Impact
of near-future ocean acidification on echinoderms.
Ecotoxicology 19:449–62.

Eggert, A., Peters, A. F. & Küpper, F. C. 2010. The potential
impacts of climate change on endophyte infections in kelp
sporophytes. In Israel, A., Einav, R. & Seckbach, J. [Eds.]
Seaweeds and Their Role in Globally Changing Environments.
Springer, New York, pp. 139–54.

Eggert, A. 2012. Seaweed responses to temperature. In Wiencke,
C. & Bischof, K. [Eds.] Seaweed Biology. Springer-Verlag,
Berlin, Germany, pp. 47–66.

Elfwing, T. & Tedengren, M. 2002. Effects of copper and reduced
salinity on grazing activity and macroalgae production: a
short-term study on a mollusc grazer, Trochus maculatus and
two species of macroalgae in the inner Gulf of Thailand.
Mar. Biol. 140:913–9.

Fabricius, K. E., Langdon, C., Uthicke, S., Humphrey, C.,
Noonan, S., De’ath, G., Okazaki, R., Muehllehner, N., Glas,
M. S. & Lough, J. M. 2011. Losers and winners in coral reefs
acclimatized to elevated carbon dioxide concentrations.
Nature Clim. Change 1:165–9.

Fain, S. R. & Murray, S. N. 1982. Effects of light and temperature
on net photosynthesis and dark respiration of gametophytes
and embryonic sporophytes of Macrocystis pyrifera. J. Phycol.
18:92–8.

Falkenberg, L. J., Russell, B. D. & Connell, S. D. 2012. Stability of
strong species interactions resist the synergistic effects of
local and global pollution in kelp forests. PLoS ONE 7:
e33841.

Feely, R. A., Sabine, C. L., Hernandez-Ayon, J. M., Ianson, D. &
Hales, B. 2008. Evidence for upwelling of corrosive “acidified”
water onto the continental shelf. Science 320:1490–2.

Feely, R. A., Sabine, C. L., Lee, K., Berelson, W., Kleypas, J.,
Fabry, V. J. & Millero, F. J. 2004. Impact of anthropogenic
CO2 on the CaCO3 system in the oceans. Science 305:362–6.

Folke, C., Carpenter, S., Walker, B., Scheffer, M., Elmqvist, T.,
Gunderson, L. & Holling, C. S. 2004. Regime shifts,
resilience, and biodiversity in ecosystem management. Annu.
Rev. Ecol. Syst. 35:557–81.

Forrest, B. M. & Blakemore, K. A. 2006. Evaluation of treatments
to reduce the spread of a marine plant pest with aquaculture
transfers. Aquaculture 257:333–45.

Gachon, C. M. M., Sime-Ngando, T., Strittmatter, M.,
Chambouvet, A. & Kim, G. H. 2010. Algal diseases: spotlight
on a black box. Trends Plant Sci. 15:633–40.

Gao, K., Aruga, Y., Asada, K., Ishihara, T., Akano, T. & Kiyohara,
M. 1993. Calcification in the articulated coralline alga
Corallina pilulifera, with special reference to the effect of
elevated CO2 concentration. Mar. Biol. 117:129–32.

Gao, K. S., Ji, Y. & Aruga, Y. 1999. Relationship of CO2
concentrations to photosynthesis of intertidal macroalgae
during emersion. Hydrobiologia 399:355–9.

Gao, K. S. & Zheng, Y. Q. 2010. Combined effects of ocean
acidification and solar UV radiation on photosynthesis,
growth, pigmentation and calcification of the coralline alga
Corallina sessilis (Rhodophyta). Global Change Biol. 16:
2388–98.

Giesecke, T., Miller, P. A., Sykes, M. T., Ojala, A. E. K., Seppa, H.
& Bradshaw, R. H. W. 2010. The effect of past changes in
inter-annual temperature variability on tree distribution
limits. J. Biogeogr. 37:1394–405.

Graham, M. H. 2004. Effects of local deforestation on the
diversity and structure of Southern California giant kelp
forest food webs. Ecosystems 7:341–57.

Graham, M. H., Harrold, C., Lisin, S., Light, K., Watanabe, J. M.
& Foster, M. S. 1997. Population dynamics of giant kelp
Macrocystis pyrifera along a wave exposure gradient. Mar. Ecol.
Prog. Ser. 148:269–79.

Graham, M. H., Kinlan, B. P., Druehl, L. D., Garske, L. E. &
Banks, S. 2007. Deep-water refugia as potential hotspots of
tropical marine diversity and productivity. Proc. Nat. Acad.
Sci. U.S.A. 104:16576–80.

Hall-Spencer, J. M., Rodolfo-Metalpa, R., Martin, S., Ransome, E.,
Fine, M., Turner, S. M., Rowley, S. J., Tedesco, D. & Buia, M. C.
2008. Volcanic carbon dioxide vents show ecosystem effects of
ocean acidification. Nature 454:96–9.

Harley, C. D. G. 2003. Abiotic stress and herbivory interact to set
range limits across a two-dimensional stress gradient. Ecology
84:1477–88.

Harley, C. D. G. 2011. Climate change, keystone predation, and
biodiversity loss. Science 334:1124–7.

Harley, C. D. G., Hughes, A. R., Hultgren, K. M., Miner, B. G.,
Sorte, C. J. B., Thornber, C. S., Rodriguez, L. F., Tomanek,
L. & Williams, S. L. 2006. The impacts of climate change in
coastal marine systems. Ecol. Lett. 9:228–41.

Harley, C. D. G. & Paine, R. T. 2009. Contingencies and compounded
rare perturbations dictate sudden distributional shifts during
periods of gradual climate change. Proc. Nat. Acad. Sci. U.S.A.
106:11172–6.

Harrold, C., Light, K. & Lisin, S. 1998. Organic enrichment of
submarine-canyon and continental-shelf benthic communities
by macroalgal drift imported from nearshore kelp forests.
Limnol. Oceanogr. 43:669–78.

Hart, M. W. & Scheibling, R. E. 1988. Heat waves, baby booms,
and the destruction of kelp beds by sea urchins. Mar. Biol.
99:167–76.

Hawkins, S. J., Moore, P. J., Burrows, M. T., Poloczanska, E.,
Mieszkowska, N., Herbert, R. J. H., Jenkins, S. R., Thompson,
R. C., Genner, M. J. & Southward, A. J. 2008. Complex
interactions in a rapidly changing world: responses of rocky
shore communities to recent climate change. Climate Res.
37:123–33.

CLIMATE CHANGE AND SEAWEEDS 1075

M
IN
IR
E
V
IE
W

Hawkins, S. J., Sugden, H. E., Mieszkowska, N., Moore, P. J.,
Poloczanska, E., Leaper, R., Herbert, R. J. H. et al. 2009.
Consequences of climate-driven biodiversity changes for
ecosystem functioning of North European rocky shores. Mar.
Ecol. Prog. Ser. 396:245–59.

Hay, M. E., Kappel, Q. E. & Fenical, W. 1994. Synergisms in plant
defenses against herbivores – interactions of chemistry,
calcification, and plant quality. Ecology 75:1714–26.

Helmuth, B., Broitman, B. R., Blanchette, C. A., Gilman, S., Halpin,
P., Harley, C. D. G., O’Donnell, M. J., Hofmann, G. E.,
Menge, B. & Strickland, D. 2006. Mosaic patterns of thermal
stress in the rocky intertidal zone: implications for climate
change. Ecol. Monogr. 76:461–79.

Hemmi, A. & Jormalainen, V. 2002. Nutrient enhancement
increases performance of a marine herbivore via quality of
its food alga. Ecology 83:1052–64.

Hepburn, C. D., Pritchard, D. W., Cornwall, C. E., McLeod, R. J.,
Beardall, J., Raven, J. A. & Hurd, C. L. 2011. Diversity of
carbon use strategies in a kelp forest community: implications
for a high CO2 ocean. Global Change Biol. 17:2488–97.

Hernandez, J. C., Clemente, S., Girard, D., Perez-Ruzafa, A. &
Brito, A. 2010. Effect of temperature on settlement and
postsettlement survival in a barrens-forming sea urchin. Mar.
Ecol. Prog. Ser. 413:69–80.

Hoegh-Guldberg, O., Mumby, P. J., Hooten, A. J., Steneck, R. S.,
Greenfield, P., Gomez, E., Harvell, C. D. et al. 2007. Coral
reefs under rapid climate change and ocean acidification.
Science 318:1737–42.

Hoffman, J. R., Hansen, L. J. & Klinger, T. 2003. Interactions
between UV radiation and temperature limit inferences from
single-factor experiments. J. Phycol. 39:268–72.

Hönisch, B., Ridgwell, A., Schmidt, D. N., Thomas, E., Gibbs, S. J.,
Sluijs, A., Zeebe, R. et al. 2012. The geological record of
ocean acidification. Science 335:1058–63.

Hughes, T. P., Baird, A. H., Bellwood, D. R., Card, M., Connolly,
S. R., Folke, C., Grosberg, R. et al. 2003. Climate change,
human impacts, and the resilience of coral reefs. Science
301:929–33.

Hunt, L. J. H. & Denny, M. W. 2008. Desiccation protection and
disruption: a trade-off for an intertidal marine alga. J. Phycol.
44:1164–70.

IPCC 2007. Climate Change 2007: The Physical Science Basis.
Contribution of Working Group I to the Fourth Assessment
Report of the Intergovernmental Panel on Climate Change.
Cambridge University Press, Cambridge, UK.

Johnson, V. R., Russell, B. D., Fabricius, K. E., Brownlee, C. &
Hall-Spencer, J. M. 2012. Temperate and tropical brown
macroalae thrive, despite decalcification, along natural CO2
gradients. Global Change Biol. 18:2792–803.

Jokiel, P. L., Rodgers, K. S., Kuffner, I. B., Andersson, A. J., Cox,
E. F. & Mackenzie, F. T. 2008. Ocean acidification and
calcifying reef organisms: a mesocosm investigation. Coral
Reefs 27:473–83.

Kavanaugh, M. T., Nielsen, K. J., Chan, F. T., Menge, B. A.,
Letelier, R. M. & Goodrich, L. M. 2009. Experimental
assessment of the effects of shade on an intertidal kelp: do
phytoplankton blooms inhibit growth of open-coast
macroalgae? Limnol. Oceanogr. 54:276–88.

Kendall, M. A., Burrows, M. T., Southward, A. J. & Hawkins, S. J.
2004. Predicting the effects of marine climate change on the
invertebrate prey of the birds of rocky shores. Ibis 146:40–7.

Kim, E., Park, H. S., Jung, Y., Choi, D. W., Jeong, W. J., Hwang,
M. S., Park, E. J. & Gong, Y. G. 2011. Identification of the
high-temperature response genes from Porphyra seriata
(Rhodophyta) expression sequence tags and enhancement
of heat tolerance of Chlamydomonas (Chlorophyta) by
expression of the Porphyra HTR2 gene. J. Phycol. 47:821–8.

Kordas, R. L., Harley, C. D. G. & O’Connor, M. I. 2011.
Community ecology in a warming world: the influence of
temperature on interspecific interactions in marine systems.
J. Exp. Mar. Biol. Ecol. 400:218–26.

Kroeker, K. J., Kordas, R. L., Crim, R. N. & Singh, G. G. 2010.
Meta-analysis reveals negative yet variable effects of

ocean acidification on marine organisms. Ecol. Lett. 13:
1419–34.

Krumhansl, K. A., Lee, J. M. & Scheibling, R. E. 2011. Grazing
damage and encrustation by an invasive bryozoan reduce the
ability of kelps to withstand breakage by waves. J. Exp. Mar.
Biol. Ecol. 407:12–8.

Kübler, J. E., Johnston, A. M. & Raven, J. A. 1999. The effects of
reduced and elevated CO2 and O2 on the seaweed
Lomentaria articulata. Plant, Cell Environ. 22:1303–10.

Kuffner I. B., Andersson, A. J., Jokiel, P. L., Rodgers, K. S. &
Mackenzie, F. T. 2008. Decreased abundance of crustose
coralline algae due to ocean acidification. Nature Geosci.
1:114–7.

Ladah, L. B. & Zertuche-González, J. A. 2007. Survival of
microscopic stages of a perennial kelp (Macrocystis pyrifera)
from the center and the southern extreme of its range in
the Northern Hemisphere after exposure to simulated El
Nino stress. Mar. Biol. 152:677–86.

Largo, D. B., Fukami, K. & Nishijima, T. 1995. Occasional
pathogenic bacteria promoting ice-ice disease in the
carrageenan-producing red algae Kappaphycus alvarezii and
Eucheuma denticulatum (Solieriaceae, Gigartinales,
Rhodophyta). J. Appl. Phycol. 7:545–54.

Leonard, G. H. 2000. Latitudinal variation in species interactions:
a test in the New England rocky intertidal zone. Ecology
81:1015–30.

Lewis, S., Donkin, M. E. & Depledge, M. H. 2001. Hsp70
expression in Enteromorpha intestinalis (Chlorophyta) exposed
to environmental stressors. Aquat. Toxicol. 51:277–91.

Li, R. & Brawley, S. H. 2004. Improved survival under heat stress
in intertidal embryos (Fucus spp.) simultaneously exposed to
hypersalinity and the effect of parental thermal history. Mar.
Biol. 144:205–13.

Lima, F. P., Ribeiro, P. A., Queiroz, N., Hawkins, S. J. & Santos,
A. M. 2007. Do distributional shifts of northern and
southern species of algae match the warming pattern? Global
Change Biol. 13:2592–604.

Ling, S. D. 2008. Range expansion of a habitat-modifying species
leads to loss of taxonomic diversity: a new and impoverished
reef state. Oecologia 156:883–94.

Lobban, C. S. & Harrison, P. J. 1997. Seaweed Ecology and
Physiology. Cambridge University Press, New York.

Long, S. P., Ainsworth, E. A., Rogers, A. & Ort, D. R. 2004. Rising
atmospheric carbon dioxide: plants face the future. Ann.
Rev. Plant Biol. 55:591–628.

Lubchenco, J. & Gaines, S. D. 1981. A unified approach to
marine plant-herbivore interactions. I. Populations and
communities. Ann. Rev. Ecol. Syst. 12:405–37.

Lüning, K. 1990. Seaweeds – Their Environment, Biogeography, and
Ecophysiology. John Wiley & Sons Inc., New York.

Lüning, K. & Neushul, M. 1978. Light and temperature
demands for growth and reproduction of Laminarian
gametophytes in southern and central California. Mar. Biol.
45:297–309.

Martin, S. & Gattuso, J. P. 2009. Response of Mediterranean
coralline algae to ocean acidification and elevated
temperature. Global Change Biol. 15:2089–100.

Méléder, V., Populus, J., Guillaumont, B., Perrot, T. & Mouquet,
P. 2010. Predictive modelling of seabed habitats: case study
of subtidal kelp forests on the coast of Brittany. France. Mar.
Biol. 157:1525–41.

Mieszkowska, N., Kendall, M. A., Hawkins, S. J., Leaper, R.,
Williamson, P., Hardman-Mountford, N. J. & Southward, A.
J. 2006. Changes in the range of some common rocky shore
species in Britain – a response to climate change?
Hydrobiologia 555:241–51.

Miller, K. A., Aguilar-Rosas, L. E. & Pedroche, F. F. 2011. A
review of non-native seaweeds from California, USA and Baja
California, Mexico. Hidrobiologica 21:365–79.

Mumby, P. J. & Harborne, A. R. 2010. Marine reserves enhance
the recovery of corals on caribbean reefs. PLoS ONE 5:e8657.

Muñoz, V., Hernández-González, M. C., Buschmann, A. H.,
Graham, M. H. & Vásquez, J. A. 2004. Variability in per

1076 CHRISTOPHER D. G. HARLEY ET AL.

M
IN
IR
E
V
IE
W

capita oogonia and sporophyte production from giant kelp
gametophytes (Macrocystis pyrifera, Phaeophyceae). Rev. Chil.
Hist. Nat. 77:639–47.

Norderhaug, K. N., Christie, H., Fossa, J. H. & Fredriksen, S.
2005. Fish-macrofauna interactions in a kelp (Laminaria
hyperborea) forest. J. Mar.Biol. Assoc. UK 85:1279–86.

O’Connor, M. I. 2009. Warming strengthens an herbivore-plant
interaction. Ecology 90:388–98.

O’Connor, M. I., Selig, E. R., Pinsky, M. L. & Altermatt, F. 2012.
Toward a conceptual synthesis for climate change responses.
Global Ecol. Biogeogr. 21:693–703.

O’Leary, J. K. & McClanahan, T. R. 2010. Trophic cascades result
in large-scale coralline algae loss through differential grazer
effects. Ecology 91:3584–97.

Padilla-Gamino, J. L. & Carpenter, R. C. 2007a. Seasonal
acclimatization of Asparagopsis taxiformis (Rhodophyta) from
different biogeographic regions. Limnol. Oceanogr. 52:833–42.

Padilla-Gamino, J. L. & Carpenter, R. C. 2007b. Thermal
ecophysiology of Laurencia pacifica and Laurencia nidifica
(Ceramiales, Rhodophyta) from tropical and warm-
temperate regions. J. Phycol. 43:686–92.

Parmesan, C. & Yohe, G. 2003. A globally coherent fingerprint of
climate change impacts across natural systems. Nature 421:
37–42.

Pearson, P. N. & Palmer, M. R. 2000. Atmospheric carbon
dioxide concentrations over the past 60 million years. Nature
406:695–9.

Pereyra, R. T., Bergstrom, L., Kautsky, L. & Johannesson, K. 2009.
Rapid speciation in a newly opened postglacial marine
environment, the Baltic Sea. BMC Evol. Biol. 9:70.

Perry, A. L., Low, P. J., Ellis, J. R. & Reynolds, J. D. 2005. Climate
change and distribution shifts in marine fishes. Science
308:1912–5.

Pörtner, H. O. & Farrell, A. P. 2008. Physiology and climate
change. Science 322:690–2.

Porzio, L., Buia, M. C. & Hall-Spencer, J. M. 2011. Effects of
ocean acidification on macroalgal communities. J. Exp. Mar.
Biol. Ecol. 400:278–87.

Price, N. N., Hamilton, S. L., Tootell, J. S. & Smith, J. E. 2011.
Species-specific consequences of ocean acidification for the
calcareous tropical green algae Halimeda. Mar. Ecol. Prog. Ser.
440:67–78.

Rabalais, N. N., Turner, R. E., Diaz, R. J. & Justic, D. 2009. Global
change and eutrophication of coastal waters. ICES J. Mar. Sci.
66:1528–37.

Raven, J. A. & Geider, R. J. 1988. Temperature and algal growth.
New Phytol. 110:441–61.

Raven, J. A., Giordano, M., Beardall, J. & Maberly, S. C. 2012.
Algal evolution in relation to atmospheric CO2: carboxylases,
carbon-concentrating mechanisms and carbon oxidation
cycles. Phil. Trans. R. Soc. B: Biol. Sci. 367:493–507.

Reuter, K. E., Lotterhos, K. E., Crim, R. N., Thompson, C. A. &
Harley, C. D. G. 2011. Elevated pCO2 increases sperm
limitation and risk of polyspermy in the red sea urchin
Strongylocentrotus franciscanus. Global Change Biol. 17:163–71.

Roleda, M. Y., Boyd, P. W. & Hurd, C. L. 2012a. Before
ocean acidification: calcifier chemistry lessons. J. Phycol.
48:840–3.

Roleda, M. Y., Morris, J. N., McGraw, C. M. & Hurd, C. L. 2012b.
Ocean acidification and seaweed reproduction: increased
CO2 ameliorates the negative effect of lowered pH on
meiospore germination in the giant kelp Macrocystis pyrifera
(Laminariales, Phaeophyceae). Global Change Biol. 18:854–64.

Rönnbäck, P., Kautsky, N., Pihl, L., Troell, M., Söderqvist, T. &
Wennhage, H. 2007. Ecosystem goods and services from
Swedish coastal habitats: identification, valuation, and
implications of ecosystem shifts. Ambio 36:534–44.

Rothäusler, E., Gomez, I., Hinojosa, I. A., Karsten, U., Tala, F. &
Thiel, M. 2009. Effect of temperature and grazing on growth
and reproduction of floating Macrocystis spp. (Phaeophyceae)
along a latitudinal gradient. J. Phycol. 45:547–59.

Rothäusler, E., Gomez, I., Karsten, U., Tala, F. & Thiel, M. 2011.
Physiological acclimation of floating Macrocystis pyrifera to

temperature and irradiance ensures long-term persistence at
the sea surface at mid-latitudes. J. Exp. Mar. Biol. Ecol. 405:
33–41.

Russell, B. D., Passarelli, C. A. & Connell, S. D. 2011. Forecasted
CO2 modifies the influence of light in shaping subtidal
habitat. J. Phycol. 47:744–52.

Russell, B. D., Thompson, J. A. I., Falkenberg, L. J. & Connell, S. D.
2009. Synergistic effects of climate change and local stressors:
CO2 and nutrient-driven change in subtidal rocky habitats.
Global Change Biol. 15:2153–62.

Saunders, M. I., Metaxas, A. & Filgueira, R. 2010. Implications of
warming temperatures for population outbreaks of a
nonindigenous species (Membranipora membranacea, Bryozoa)
in rocky subtidal ecosystems. Limnol. Oceanogr. 55:1627–42.

Scheffer, M., Carpenter, S., Foley, J. A., Folke, C. & Walker, B.
2001. Catastrophic shifts in ecosystems. Nature 413:591–6.

Scheibling, R. E. & Gagnon, P. 2009. Temperature-mediated
outbreak dynamics of the invasive bryozoan Membranipora
membranacea in Nova Scotian kelp beds. Mar. Ecol. Prog. Ser.
390:1–13.

Schiel, D. R. & Foster, M. S. 2006. The Population Biology of Large
Brown Seaweeds: Ecological Consequences of Multiphase Life
Histories in Dynamic Coastal Environments. Annu. Rev. Ecol.
Evol. Syst., Annual Reviews, Palo Alto, California, pp. 343–72.

Schiel, D. R., Steinbeck, J. R. & Foster, M. S. 2004. Ten years of
induced ocean warming causes comprehensive changes in
marine benthic communities. Ecology 85:1833–9.

Schonbeck, M. W. & Norton, T. A. 1980. Factors controlling the
lower limits of fucoid algae on the shore. J. Exp. Mar. Biol.
Ecol. 43:131–50.

Serisawa, Y., Imoto, Z., Ishikawa, T. & Ohno, M. 2004. Decline of
the Ecklonia cava population associated with increased
seawater temperatures in Tosa Bay, southern Japan. Fish. Sci.
70:189–91.

Seymour, R. J., Tegner, M. J., Dayton, P. K. & Parnell, P. E. 1989.
Storm wave-induced mortality of giant kelp, Macrocystis
pyrifera, in southern California. Estuar. Coast. Shelf Sci. 28:
277–92.

Sousa, W. P. 1979. Disturbance in marine intertidal boulder
fields: the nonequilibrium maintenance of species diversity.
Ecology 60:1225–39.

Southworth, J., Randolph, J. C., Habeck, M., Doering, O. C.,
Pfeifer, R. A., Rao, D. G. & Johnston, J. J. 2000.
Consequences of future climate change and changing
climate variability on maize yields in the midwestern United
States. Agric. Ecosyst. Environ. 82:139–58.

Speidel, M., Harley, C. D. G. & Wonham, M. J. 2001. Recovery of
the brown alga Fucus gardneri following a range of removal
intensities. Aq. Bot. 71:273–80.

Steen, H. 2004a. Effects of reduced salinity on reproduction and
germling development in Sargassum muticum (Phaeophyceae,
Fucales). Eur. J. Phycol. 39:293–9.

Steen, H. 2004b. Interspecific competition between Enteromorpha
(Ulvales: Chlorophyceae) and Fucus (Fucales: Phaeophyceae)
germlings: effects of nutrient concentration, temperature,
and settlement density. Mar. Ecol. Prog. Ser. 278:89–101.

Sunday, J. M., Crim, R. N., Harley, C. D. G. & Hart, M. W. 2011.
Quantifying rates of evolutionary adaptation in response to
ocean acidification. PLoS ONE 6:e22881.

Swanson, A. K. & Fox, C. H. 2007. Altered kelp (Laminariales)
phlorotannins and growth under elevated carbon dioxide
and ultraviolet-B treatments can influence associated
intertidal food webs. Global Change Biol. 13:1696–709.

Taylor, D. I. & Schiel, D. R. 2010. Algal populations controlled by
fish herbivory across a wave exposure gradient on southern
temperate shores. Ecology 91:201–11.

Thom, R. M. 1996. CO2-enrichment effects on eelgrass (Zostera
marina L) and bull kelp (Nereocystis luetkeana (Mert) P & R).
Water Air Soil Poll. 88:383–91.

Thompson, R. C., Wilson, B. J., Tobin, M. L., Hill, A. S. &
Hawkins, S. J. 1996. Biologically generated habitat provision
and diversity of rocky shore organisms at a hierarchy of
spatial scales. J. Exp. Mar. Biol. Ecol. 202:73–84.

CLIMATE CHANGE AND SEAWEEDS 1077

M
IN
IR
E
V
IE
W

Van Alstyne, K. L., Pelletreau, K. N. & Kirby, A. 2009. Nutritional
preferences override chemical defenses in determining food
choice by a generalist herbivore, Littorina sitkana. J. Exp. Mar.
Biol. Ecol. 379:85–91.

Vaselli, S., Bertocci, I., Maggi, E. & Benedetti-Cecchi, L. 2008.
Assessing the consequences of sea level rise: effects of
changes in the slope of the substratum on sessile
assemblages of rocky seashores. Mar. Ecol. Prog. Ser. 368:9–22.

Vayda, M. E. & Yuan, M. L. 1994. The heat-shock response of an
Antarctic alga is evident at 5°C. Plant Mol. Biol. 24:229–33.

Vermeij, M. J. A., van Moorselaar, I., Engelhard, S., Hornlein, C.,
Vonk, S. M. & Visser, P. M. 2010. The effects of nutrient
enrichment and herbivore abundance on the ability of turf
algae to overgrow coral in the Caribbean. PLoS ONE 5:
e14312.

Vinueza, L. R., Branch, G. M., Branch, M. L. & Bustamante, R. H.
2006. Top-down herbivory and bottom-up El Niño effects on
Galapagos rocky-shore communities. Ecol. Monogr. 76:111–31.

Wahl, M., Jormalainen, V., Eriksson, B. K., Coyer, J. A., Molis, M.,
Schubert, H., Dethier, M. et al. 2011. Stress ecology in Fucus:
abiotic, biotic, and genetic interactions. In Lesser, M. [Ed.]
Advances in Marine Biology, Vol 59. Elsevier Academic Press
Inc, San Diego, California, pp. 37–105.

Wang, M. Y., Overland, J. E. & Bond, N. A. 2010. Climate projections
for selected large marine ecosystems. J. Mar. Syst. 79:258–66.

Weinberger, F., Rohde, S., Oschmann, Y., Shahnaz, L., Dobretsov, S.
& Wahl, M. 2011. Effects of limitation stress and of disruptive
stress on induced antigrazing defense in the bladder wrack
Fucus vesiculosus. Mar. Ecol. Prog. Ser. 427:83–94.

Wernberg, T., Russell, B. D., Thomsen, M. S., Gurgel, C. F. D.,
Bradshaw, C. J. A., Poloczanska, E. S. & Connell, S. D. 2011a.

Seaweed communities in retreat from ocean warming.
Current Biol. 21:1828–32.

Wernberg, T., Smale, D. A. & Thomsen, M. S. 2012. A decade of
climate change experiments on marine organisms: procedures,
patterns and problems. Global Change Biol. 18:1491–8.

Wernberg, T., Thomsen, M. S., Tuya, F. & Kendrick, G. A. 2011b.
Biogenic habitat structure of seaweeds change along a
latitudinal gradient in ocean temperature. J. Exp. Mar. Bio.
Ecol. 400:264–71.

Wernberg, T., Thomsen, M. S., Tuya, F., Kendrick, G. A., Staehr,
P. A. & Toohey, B. D. 2010. Decreasing resilience of kelp
beds along a latitudinal temperature gradient: potential
implications for a warmer future. Ecol. Lett. 13:685–94.

Williams, S. L. & Dethier, M. N. 2005. High and dry: variation in
net photosynthesis of the intertidal seaweed Fucus gardneri.
Ecology 86:2373–9.

Wootton, J. T., Pfister, C. A. & Forester, J. D. 2008. Dynamic
patterns and ecological impacts of declining ocean pH in a
high-resolution multi-year dataset. Proc. Nat. Acad. Sci. U.S.A.
105:18848–53.

Xu, Z. G., Zou, D. H. & Gao, K. S. 2010. Effects of elevated CO2 and
phosphorus supply on growth, photosynthesis and nutrient
uptake in the marine macroalga Gracilaria lemaneiformis
(Rhodophyta). Bot. Mar. 53:123–9.

Zacharioudaki, A., Pan, S. Q., Simmonds, D., Magar, V. & Reeve,
D. E. 2011. Future wave climate over the west-European shelf
seas. Ocean Dynam. 61:807–27.

Zou, D. & Gao, K. 2002. Effects of desiccation and CO2
concentrations on emersed photosynthesis in Porphyra
haitanensis (Bangiales, Rhodophyta), a species farmed in
China. Eur. J. Phycol. 37:587–92.

1078 CHRISTOPHER D. G. HARLEY ET AL.

M
IN
IR
E
V
IE
W

High summer temperatures amplify functional differences between
coral- and algae-dominated reef

communities

FLORIAN ROTH ,1,2,3,10 NILS RäDECKER ,1,4,5 SUSANA CARVALHO ,1 CARLOS M. DUARTE ,1,6 VINCENT
SADERNE ,1 ANDREA ANTON ,1,6 LUIS SILVA ,1 MARIA LI. CALLEJA ,1,7 XOSÉ ANXELU G. MORÁN ,1

CHRISTIAN R. VOOLSTRA ,1,4 BENJAMIN KüRTEN ,1,8 BURTON H. JONES ,1 AND CHRISTIAN WILD 9

1Red Sea Research Center, King Abdullah University of Science and Technology (KAUST), Thuwal 23955 Saudi Arabia
2Baltic Sea Centre, Stockholm University, Stockholm 10691 Sweden

3Faculty of Biological and Environmental Sciences, Tvärminne Zoological Station, University of Helsinki, Helsinki 00014 Finland
4Department of Biology, University of Konstanz, Konstanz 78457 Germany

5Laboratory for Biological Geochemistry, School of Architecture, Civil and Environmental Engineering, École Polytechnique Fédérale
de Lausanne (EPFL), Lausanne 1015 Switzerland

6Computational Biology Research Center, King Abdullah University of Science and Technology (KAUST), Thuwal 23955 Saudi
Arabia

7Department of Climate Geochemistry, Max Planck Institute for Chemistry (MPIC), Mainz 55128 Germany
8Project Management Jülich, Jülich Research Centre GmbH, Rostock 52425 Germany

9Marine Ecology, Faculty of Biology and Chemistry, University of Bremen, Bremen 28359 Germany

Citation: Roth, F., N. Rädecker, S. Carvalho, C. M. Duarte, V. Saderne, A. Anton, L. Silva, M. L. L. Call-
eja, X. A. G. Morán, C. R. Voolstra, B. Kürten, B. H. Jones, and C. Wild. 2021. High summer tempera-
tures amplify functional differences between coral- and algae-dominated reef communities. Ecology 10

2

(2):e03226. 10.1002/ecy.3226

Abstract. Shifts from coral to algal dominance are expected to increase in tropical coral
reefs as a result of anthropogenic disturbances. The consequences for key ecosystem functions
such as primary productivity, calcification, and nutrient recycling are poorly understood, par-
ticularly under changing environmental conditions. We used a novel in situ incubation
approach to compare functions of coral- and algae-dominated communities in the central Red
Sea bimonthly over an entire year. In situ gross and net community primary productivity, calci-
fication, dissolved organic carbon fluxes, dissolved inorganic nitrogen fluxes, and their respec-
tive activation energies were quantified to describe the effects of seasonal changes. Overall,
coral-dominated communities exhibited 30% lower net productivity and 10 times higher calcifi-
cation than algae-dominated communities. Estimated activation energies indicated a higher
thermal sensitivity of coral-dominated communities. In these communities, net productivity
and calcification were negatively correlated with temperature (>40% and >65% reduction,
respectively, with +5°C increase from winter to summer), whereas carbon losses via respiration
and dissolved organic carbon release more than doubled at higher temperatures. In contrast,
algae-dominated communities doubled net productivity in summer, while calcification and dis-
solved organic carbon fluxes were unaffected. These results suggest pronounced changes in
community functioning associated with coral-algal phase shifts. Algae-dominated communities
may outcompete coral-dominated communities because of their higher productivity and car-
bon retention to support fast biomass accumulation while compromising the formation of
important reef framework structures. Higher temperatures likely amplify these functional dif-
ferences, indicating a high vulnerability of ecosystem functions of coral-dominated communi-
ties to temperatures even below coral bleaching thresholds. Our results suggest that ocean
warming may not only cause but also amplify coral–algal phase shifts in coral reefs.

Key words: activation energy; biogeochemical cycling; climate change; community budget; ecosystem
functioning; regime shifts.

INTRODUCTION

Community shifts and the ongoing loss of biodiversity
(Brondizio et al. 2019) are altering the productivity and
biogeochemistry of many ecosystems globally

(Middleton and Grace 2004, Hooper et al. 2012, Naeem
et al. 2012). These changes compound with local and
global environmental perturbations, which can acceler-
ate the alteration of essential ecosystem processes (Bal-
vanera et al. 2006, Stachowicz et al. 2007). Thermal
stress caused by climate change is, thereby, likely to exhi-
bit the most substantial impact (Stillman 2019).
Tropical coral reefs are hotspots of biodiversity that

provide various ecosystem services that are supported by

Manuscript received 17 March 2020; revised 6 July 2020;
accepted 24 August 2020. Corresponding Editor: Richard B.
Aronson.

10 E-mail: florian.roth@su.se

Article e03226; page 1

Ecology, 102(2), 2021, e03226
© 2020 The Authors. Ecology published by Wiley Periodicals LLC on behalf of Ecological Society of America
This is an open access article under the terms of the Creative Commons Attribution License, which permits use, distribution and reproduction in any
medium, provided the original work is properly cited.

https://orcid.org/0000-0003-4004-5863

https://orcid.org/0000-0003-4004-5863

https://orcid.org/0000-0003-4004-5863

https://orcid.org/0000-0002-2387-8567

https://orcid.org/0000-0002-2387-8567

https://orcid.org/0000-0002-2387-8567

https://orcid.org/0000-0003-1300-1953

https://orcid.org/0000-0003-1300-1953

https://orcid.org/0000-0003-1300-1953

https://orcid.org/0000-0002-1213-1361

https://orcid.org/0000-0002-1213-1361

https://orcid.org/0000-0002-1213-1361

https://orcid.org/0000-0003-3968-2718

https://orcid.org/0000-0003-3968-2718

https://orcid.org/0000-0003-3968-2718

https://orcid.org/0000-0002-4104-2966

https://orcid.org/0000-0002-4104-2966

https://orcid.org/0000-0002-4104-2966

https://orcid.org/0000-0001-8434-6559

https://orcid.org/0000-0001-8434-6559

https://orcid.org/0000-0001-8434-6559

https://orcid.org/0000-0002-5992-2013

https://orcid.org/0000-0002-5992-2013

https://orcid.org/0000-0002-5992-2013

https://orcid.org/0000-0002-9823-5339

https://orcid.org/0000-0002-9823-5339

https://orcid.org/0000-0002-9823-5339

https://orcid.org/0000-0003-4555-379

5

https://orcid.org/0000-0003-4555-3795

https://orcid.org/0000-0003-4555-3795

https://orcid.org/0000-0003-0328-7847

https://orcid.org/0000-0003-0328-7847

https://orcid.org/0000-0003-0328-7847

https://orcid.org/0000-0002-9599-1593

https://orcid.org/0000-0002-9599-1593

https://orcid.org/0000-0002-9599-1593

https://orcid.org/0000-0001-9637-6536

https://orcid.org/0000-0001-9637-6536

https://orcid.org/0000-0001-9637-6536

info:doi/10.1002/ecy.3226

mailto:

http://creativecommons.org/licenses/by/4.0/

http://crossmark.crossref.org/dialog/?doi=10.1002%2Fecy.3226&domain=pdf&date_stamp=2020-12-27

one or more metabolic or biogeochemical functions
(e.g., primary production, calcification, organic matter
fluxes, and nutrient cycling; Moberg and Folke 1999).
Many of these processes are primarily driven by sclerac-
tinian corals, the “ecosystem engineers” of tropical reefs
(Wild et al. 2011). However, the combination of global
and local anthropogenic stressors has caused extensive
coral mortality and subsequent shifts from complex
coral-dominated communities to simplified communities
with a predominance of filamentous turf- and macroal-
gae in many reefs around the world (Done 1992, Bell-
wood et al. 2004, Hughes et al. 2007, Graham et al.
2015). Although coral–algal phase shifts are increasingly
observed globally, the consequences for reef ecosystem
functions such as productivity, calcification, and nutri-
ent cycling are poorly understood. Laboratory and
mesocosm studies indicate that reef algae, particularly
the widespread filamentous turfs, are metabolically very
different from corals, and generally display significantly
higher primary production rates (Rix et al. 2015, Cardini
et al. 2016). At the same time, the fraction of the photo-
synthetically fixed carbon (C) being exuded into the
environment is generally more labile (Nelson et al.
2013). At the community level, these differences may
result in changes in the carbonate chemistry of seawater
(McMahon et al. 2013, Bernstein et al. 2016), disrupted
trophic structures (Johnson et al. 1995, Hempson et al.
2018), or increased microbial loads on algae-dominated
reefs worldwide (Jessen et al. 2013, Haas et al. 2016).
Divergent responses to changing environmental condi-

tions may amplify ecosystem functions of corals and
algae differently. As such, changing temperature regimes
and recurrent heatwaves, which are increasing in fre-
quency and magnitude (Frölicher et al. 2018, Oliver
et al. 2018), can have detrimental effects on tropical
coral reef taxa (Lough et al. 2018, Hughes et al. 2019).
In corals, sublethal heat stress during summer can

compromise primary production and calcification (Rey-
naud et al. 2003, Anthony et al. 2008), thereby altering
the release of organic and inorganic products (Niggl
et al. 2009, Piggot et al. 2009). In contrast, benthic turf-
and macroalgae may be less sensitive to heat (Koch et al.
2013), showing increased productivity and net growth
with rising temperature (Bender et al. 2014). Likewise,
temperature-related productivity optima and mortality
thresholds of algae are often well above those of corals
(Anton et al. 2020). Similarly, the abundance of reef
algae can increase seasonally, especially during the sum-
mer months (Lirman and Biber 2000, Diaz-Pulido and
Garzón-Ferreira 2002, Ateweberhan et al. 2006).
However, few studies investigated the effects of coral–-

algal phase shifts on community metabolism, particu-
larly in situ. This paucity of information probably
reflects the logistical challenges of quantifying the func-
tions of structurally complex communities in their natu-
ral environment (Roth et al. 2019). Currently, most data
describing ecosystem functions are derived from labora-
tory (e.g., Cardini et al. 2016) and mesocosm (e.g.,

Langdon et al. 2003, Bellworthy and Fine 2018,
Edmunds et al. 2020) studies using either single organ-
isms or simplified reconstructed communities to predict
in situ changes at the community scale. Although these
approaches provide valuable mechanistic insights and
permit a tight control of environmental conditions dur-
ing the experiments, they can only approximate natural
conditions. However, primary production, calcification,
and organic matter recycling critically depend on local
environmental conditions, biodiversity, and system
heterogeneity (Baird et al. 2007). In addition, large parts
of the energy and nutrient pool are remineralized by
microbial communities or cryptic fauna within the reef
matrix (Richter and Wunsch 1999, de Jongh and Van
Duyl 2004, Maldonado et al. 2012), all of which are gen-
erally not considered in ex situ experimental setups.
Concordantly, Roth et al. (2019) highlighted in a com-
parison between laboratory-based single-organism and
in situ incubations that ex situ measurements that are
scaled up to average-constructed communities can over-
estimate community-wide net primary production and
underestimate respiration and gross photosynthesis by
20–90%. Hence, laboratory experiments can only pro-
vide a glimpse of the complex environmental dynamics
(e.g., seasonality) that shape the ecological processes of
reef communities (Damgaard 2019).
To overcome these experimental constraints, we used

a novel in situ approach that allowed the quantification
of major metabolic and biogeochemical pathways (Roth
et al. 2019) of co-occurring natural coral- and algae-
dominated reef communities in the Red Sea. With a total
of 112 light and dark in situ incubations, we measured
rates of community production (i.e., net community pro-
duction [NCP], community respiration [CR], and gross
primary production [GPP]), net community calcification
(NCC), net dissolved organic carbon (DOC), and dis-
solved inorganic nitrogen (DIN) fluxes bimonthly for an
entire year. In addition, we quantified the thermal-de-
pendence of the functioning of benthic communities by
applying principles of the metabolic theory of ecology
(MTE; Sibly et al. 2012). We quantified the temperature
sensitivity of metabolic processes using the activation
energy (Ea), as the slope or rate of change in the rise and
falling phases of a thermal performance curve before
and after achieving the optimal temperature. Although
activation energies are commonly assessed at the organ-
ism level (Garcı́a et al. 2018, Savva et al. 2018, Anton
et al. 2020), they also provide useful insights regarding
the sensitivity of community metabolism to warming
(Follstad Shah et al. 2017, Morán et al. 2017, Padfield
et al. 2017).
Thus, we (1) directly compare the magnitudes and

directions of key functions of coral-dominated and
phase-shifted algae-dominated reef communities, (2

)

derive their functional responses to environmental
changes induced by seasonality, and (3) describe their
thermal sensitivity to seasonally variable temperature
changes.

Article e03226; page 2 FLORIAN ROTH ET AL. Ecology, Vol. 102, No. 2

MATERIALS AND METHODS

Study site and environmental conditions

This experiment was carried out at Abu Shosha reef
located in the central Red Sea on the west coast of
Saudi Arabia (22°18’16.3’’ N; 39°02’57.7’’ E) from
January 2017 until January 2018. Key environmental
variables were monitored at the sampling site and were
previously reported in Roth et al. (2018). Water temper-
ature was measured continuously (logging interval =
30 min) for the whole study period with Onset HOBO
temperature/light data loggers (accuracy: �0.2°C)
deployed at the seafloor. Salinity was measured at each
day of sampling with a WTW TetraCon® conductivity
cell (accuracy: �0.5% of value). Light availability was
measured continuously (logging interval = 1 min) on
three full days per month with the above-mentioned
Onset HOBO data logger. Light readings were con-
verted from lux to photosynthetically active radiation
(PAR; μmol quanta�m−2�s−1; 400–700 nm wavelengths)
by intercalibration and conversion as outlined in Roth
et al. (2018), and values are presented as daytime
means. Seawater samples for the determination of dis-
solved nitrate (NO�3 ), nitrite (NO

�
2 ), ammonium

(NHþ4 ), phosphate (PO
3�
4 ), and monomeric silicate (Si

(OH)4) were taken in triplicates each month from 1 m
above the seafloor. Details for sampling and analysis
can be found in Appendix S1: Section S1. The sum of
NO�3 , NO

�
2 , and NH

þ
4 is termed “dissolved inorganic

nitrogen” (DIN) henceforth.

Benthic communities selected for in situ incubations

Abu Shosha reef is characterized by a heterogeneous
mosaic of patches of coral- and algae-dominated com-
munities. Thus, this site allows for the quantification of
the functionality of both communities under identical
environmental conditions. Nnatural coral- and algae-
dominated reef communities surrounded by sand were
haphazardly selected at the study site at 5 m water depth
within an area of 50 × 50 m. The communities had to
fulfill the following characteristics to qualify as “suitable
candidates” for later incubations: (1) Coral-dominated
communities were defined by having >40% coral cover
but <10% algae cover; (2) algae-dominated communities were defined by having >40% algae cover but <10% coral cover; (3) all communities had to fit into the incu- bation chambers (max. diameter 50 cm, max. height 39 cm). Among all suitable candidates in the study area, four coral-dominated and four algae-dominated com- munities were chosen randomly. These eight communi- ties were revisited each month of sampling. The community composition at the level of major

functional groups was assessed for each community
three times during the study period (i.e., in the begin-
ning, after 6 months, and at the end of the experiments).
Details on the assessment and statistical evaluation of

the community composition can be found in Appendix
S1: Section S1. Two-factor permutational multivariate
analysis of variance (PERMANOVA) indicated that dif-
ferences between communities grouped according to
coral and algal dominance were significant (P = 0.001;
visualization in Appendix S1: Fig. S1). As no significant
changes over time in the relative benthic cover were
detected (Appendix S1: Table S2),the benthic commu-
nity composition of each treatment was averaged over
all replicates and survey points (Fig. 1a).
Algae-dominated communities were considered

“phase-shifted,” as complex structures and the occur-
rence of coral rubble indicated that branching corals
were present previously. The co-occurrence of coral and
algal reef communities within small spatial ranges was
reported as “mosaic-dynamics” before (e.g., Edmunds
2002, Tkachenko et al. 2007) and may be explained by a
combination of local processes and historical effects,
such as previous stress events or adaptation (Done et al.
1991, Bythell et al. 2000, Edmunds 2002).
Cryptic habitats can encompass about 60–75% of the

total surface area of a reef (Richter and Wunsch 1999,
Richter et al. 2001), but organisms living in cracks and
crevices within the communities’ matrices could not be
assessed by our conventional benthic surveys. These
organisms (e.g., sponges, bryozoans, and tunicates),
however, metabolize organic matter in the order of
15–30% of the gross production of a coral reef (reviewed
in de Jongh and Van Duyl 2004), driving community res-
piration and other biogeochemical fluxes assessed in this
study. As our benthic incubation chambers jointly cap-
tured the metabolism of all members of the communities
(i.e., from the visible surface and cryptic habitats), we
refrained from assigning measured metabolic activities
to functional groups on the visible surface only, as the
inferred contribution would be highly biased. Thus,
measurements presented in this study represent commu-
nity-wide processes that include all compartments of the
reef benthos and the surrounding water.

In situ incubations and quantification of community
functions

In situ incubations with benthic chambers were per-
formed according to the protocol described in Roth
et al. (2019). In brief, chambers were constructed from
polymethyl methacrylate (PMMA) cylinders with
removable gastight lids of the same material. All cham-
bers were equipped with individual water circulation
pumps with adjustable flow control, autonomous
recording dissolved oxygen (DO), and temperature sen-
sors (HOBO U26; temperature corrected and salinity
adjusted), and two sampling ports for discrete water
samples. Incubations were carried out on three consecu-
tive days in January, March, May, July, September, and
November 2017, and in January 2018. Generally, on day
1, divers deployed four chambers on coral-dominated,
and four chambers on algae-dominated reef

February 2021 FUNCTIONING OF CORAL REEF COMMUNITIES Article e03226; page 3

communities (Fig. 1a). The chambers were positioned
carefully and left in place with open tops (no lids) until
the next morning. On the second day, incubations
started at around 09:00 a.m. by tightly securing the lids
and closing all sampling ports during natural daylight
conditions. The exact incubation start and end time was
recorded for each chamber. Incubations ran for approxi-
mately 2 h. The chambers were left in place with open
tops for a second set of incubation on the following day.
On day 3, benthic communities were incubated at “simu-
lated” darkness during the same period used for incuba-
tions the previous day. The procedure followed the same
as on day two; however, all chambers were covered with
thick black PVC covers. Any light penetration into the
chambers was prevented, as validated by control read-
ings of Onset HOBO temperature/light data loggers
within chambers. Between deployments, all materials
were rinsed with freshwater, washed with 4% hydrochlo-
ric acid (HCl), and subsequently rinsed with deionized

water for reliable water chemistry measurements that
included sensitive DOC samples.
Discrete water samples for dissolved inorganic carbon

(DIC), total alkalinity (TA), DOC, and DIN were with-
drawn from the sampling ports with acid-washed syr-
inges at the beginning and the end of each incubation
(details for the analysis of these samples can be found in
Appendix S1: Section S1). Changes in seawater chem-
istry between start and end of incubations were used to
calculate rates of NCP, CR, GPP (calculated as GPP =
NCP + |CR|), NCC, and fluxes of DOC, and DIN. All
rates and fluxes were extrapolated to incubation water
volume (in L) and normalized to incubation duration (in
h) and the planar reef area (in m2) of the enclosed ben-
thic community adapted after Roth et al. (2019).
Productivity and respiration rates (NCP and CR; in

mmol C�m−2�h−1) were calculated by changes in DIC
concentrations, taking into account calcification and
dissolution rates according to an adapted protocol by

(a)

(b)

(c)

(

(

(%)

FIG 1. Experimental setup, relative benthic cover, and key environmental variables at the study site. (a) Relative benthic cover of
functional groups in the studied coral- and algae-dominated reef communities, and exemplary pictures of the incubation chambers
on the respective substrates. Details on the community composition of each replicate and time point can be found in Appendix S1:
Section S1 and Fig. S1. (b) Photosynthetically active radiation (PAR, in μmol photons�m−2�s−1), dissolved inorganic nitrogen (DIN,
in μM), and aragonite saturation state (Ωarag) at the study site from January 2017 until January 2018. Circles represent values of dis-
crete samples for DIN and Ωarag, and daytime averages of three separate days per month for PAR; lines represent the smoothed
trend through the means. (c) Plot of seawater temperature at experimental site from January 2017 until January 2018. Each dot rep-
resents one measurement at 30-min intervals. Dashed horizontal line depicts the mean maximum annual temperature modeled for
the region from 1982 to 2015, taken from Chaidez et al. (2017). [Color figure can be viewed at wileyonlinelibrary.com]

Article e03226; page 4 FLORIAN ROTH ET AL. Ecology, Vol. 102, No. 2

www.wileyonlinelibrary.com

Albright et al. (2013). Rates of NCP and CR based on
DIC fluxes were compared to rates based on oxygen
fluxes from continuous measurements with DO sensors.
No discrepancy between C and oxygen measurements
was detected (r = 0.99, P < 0.0001, n = 112). The calcu- lated photosynthetic (1.05 � 0.02) and respiratory (0.97 � 0.02) quotients agree with those obtained by various authors elsewhere, typically ∼1 mol of oxygen produced for 1 mol C fixed, and vice versa (e.g., Gattuso et al. 1999b, Atkinson and Falter 2003). NCC (in mmol CaCO3�m−2�h−1) was calculated by

concentration differences in TA, which are primarily
caused by calcification and dissolution of CaCO3,
whereby TA is reduced (increased) by two molar equiva-
lents for every mole of CaCO3 produced (dissolved)
(Zeebe and Wolf-Gladrow 2001). Nutrients fluxes (i.e.,
NO�3 , NH

þ
4 , PO

3�
4 , and SO

2�
4 ) that cause a change in TA

unrelated to calcification and dissolution were
accounted for according to Zeebe and Wolf-Gladrow
(2001) and Wolf-Gladrow et al. (2007). DOC (in mmol
C�m−2�h−1) and DIN (in µmol N�m−2�h−1) fluxes were
calculated from concentration differences between start
and endpoints. Any temperature corrections that were
necessary for seawater chemistry calculations were
achieved by temperature readings from individual tem-
perature loggers within each chamber.
Although measurements presented in this study only

relate to small benthic communities, most studies cur-
rently work with individual reef organisms (e.g., Anton
et al. 2020) or reconstructed communities (e.g.,
Edmunds et al. 2020) to derive community functions.
Thus, the results presented here are among the best
approximations for the quantification of community-
wide (biogeochemical) ecosystem functions of
untouched, natural coral reef communities in situ (but
see Haas et al. 2013, Van Heuven et al. 2018).

Data analysis

Statistical analyses were performed using JMP©
Pro14 (SAS Institute) statistic software. Environmental
variables and response parameters from incubations
were grouped into spring (March–May), summer
(June–September), fall (October–November), and winter
(December–February) for statistical analysis. For the
seasonal comparison and to derive GPP/CR ratios,
hourly rates from light and dark incubations were used
to calculate daily net fluxes. We acknowledge that there
is a chance for a slight over- or underestimation because
of associated changes in environmental conditions (e.g.,
light) during the course of the day. Thus, to minimize
the error associated with extrapolating hourly rates, we
chose a time window for daylight incubations (from
around 09:00 a.m. to 11:00 a.m.) that is closest to day-
time average irradiation and excludes the “ramping up”
phase in the early morning hours and extreme values
that can occur during midday. As all incubations during
all sampling periods were conducted at the very same

time of the day, incubations are comparable across com-
munity types and time points.
The full seawater carbonate system parameters were

derived for each sampling period from measured salinity,
temperature, nutrients, TA, and DIC data using the
CO2SYS Microsoft Excel Macro by Pierrot et al. (2006)
and the R package Seacarb (Lavigne and Gattuso 2013)
(Appendix S1: Table S3). Environmental variables were
tested for differences with two-tailed t-tests. Response
parameters from incubations (GPP, NCP, CR, NCC,
DOC, and DIN) were assessed by linear mixed models
(LMMs) to test for differences in the respective response
parameters with ‘treatment’ (coral- vs. algae-dominated)
and ‘season’ (spring, summer, fall, and winter) as fixed
factors, and the sampling dates (date) within seasons
and the replicates of the communities (community ID)
as random factors. Tukey’s Honest Significant Differ-
ence (HSD) test was used for pairwise comparisons if
significant interactions (treatment * season) were found.
Detailed statistical results, including significant post hoc
comparisons, are presented in Appendix S1: Table S4.
The relationships between response (e.g., metabolic
functions) and explanatory variables (e.g., environmen-
tal variables) were assessed by linear regression models.
The thermal sensitivity of the metabolic processes

(GPP, NCP, CR, NCC, DOC, and DIN) was explored
by calculating the activation energy (Ea) based on
Arrhenius equations (Sibly et al. 2012) within the sea-
sonal thermal regime (25.0–32.8°C). The activation ener-
gies (Ea in eV) were estimated by fitting a linear
regression equation between the natural logarithm of the
metabolic rates and the reciprocal of temperature (1/kT),
where k is the Boltzmann’s constant (8.62 × 10−5 eV/K)
and T is the water temperature (K). To deal with obser-
vations ≤0 on log-transformed data (e.g., NCC, DOC,
and DIN rates), a constant was added (i.e., ln(rate + 1 −
min value(rate)) to shift all values above zero (Legendre
and Legendre 2012). The alternative of excluding ≤0 val-
ues was discarded because these measurements are an
important part of the biological processes under investi-
gation (Canavero et al. 2018).

RESULTS

Environmental conditions

Monthly monitored environmental variables at the
study site exhibited strong seasonal patterns (Fig. 1b,c,
Appendix S1: Table S2). The average seawater tempera-
ture ranged from 25.8 � 0.2°C in winter to
32.3 � 0.1°C in summer (Fig. 1c). Simultaneously, aver-
age daytime PAR intensities at 5 m water depth
increased from 130 � 2 μmol photons�m−2�s−1 to
465 � 14 μmol photons�m−2�s−1 (Fig. 1b). Seawater
DIN concentrations were lowest in spring and winter
(0.46 � 0.02 and 0.66 � 0.04 μM DIN, respectively)
and significantly higher in summer and fall (1.03 � 0.06
and 1.08 � 0.13 μM DIN, respectively, Fig. 1b).

February 2021 FUNCTIONING OF CORAL REEF COMMUNITIES Article e03226; page 5

Functions of coral- and algae-dominated reef communities

The rates observed along the various deployments
(Fig. 2) were used to calculate average values over the
whole study period (Table 1). Average NCP was 30%
and CR 50% lower in coral- as compared to algae-domi-
nated communities (NCP, mean � SE: 26.7 � 1.2 and
36.9 � 1.7 mmol C�m−2�h−1, respectively; CR:
−10.9 � 0.7 and −20.9 � 1.4 mmol C�m−2�h−1, respec-
tively; Fig. 2a,b). Integrated over 1 d, these differences
yielded a 40% lower GPP of coral- compared to algae-
dominated communities, with GPP/CR ratios of
2.4 � 0.1 and 2.0 � 0.2, respectively (Table 1).
NCC in the light was sixfold higher in coral- com-

pared to algae-dominated communities, averaging
7.9 � 0.7 mmol and 1.3 � 0.2 mmol CaCO3�m−2�h−1,
respectively. In the dark, coral-dominated communities
displayed an NCC of 3.0 � 0.3 mmol CaCO3�m−2�h−1,
whereas algae-dominated communities exhibited an
NCC of −0.5 � 0.2 mmol CaCO3�m−2�h−1 (represent-
ing net CaCO3 dissolution). Integrating the hourly rates
over 24 h, corals showed a 10-fold higher NCC com-
pared to algae-dominated reef communities (Table 1).
Coral-dominated communities were net sources of

DOC during both light and dark incubations, with aver-
age fluxes of 0.57 � 0.07 and 0.62 � 0.11 mmol C�m−2-
�h−1, respectively (Table 1). In contrast, algae-dominated
communities released similar amounts of DOC as corals
in the light (0.70 � 0.08 mmol C�m−2�h−1) but were net
sinks of DOC in the dark (−0.37 � 0.05 mmol C�m−2-
�h−1). When integrated over 24 h, net DOC fluxes in
coral-dominated communities were 3.5-fold higher than
those in algae-dominated communities (Table 1).
Both coral- and algae-dominated communities were net

sources of DIN (1.62 � 0.21 and 1.66 � 0.23
mmol N�m−2�d−1, respectively; Table 1), with no signifi-
cant differences between treatments. There were, however,
significant differences between light and dark incubations:
algae-dominated communities released three times more
DIN in the dark as compared to light conditions
(109.9 � 14.1 μmol N�m−2�h−1 and 28.5 � 8.7, respec-
tively). In contrast, coral-dominated communities released
DIN at consistent rates during light and in dark incuba-
tions (65.7 � 11.6 and 69.7 � 8.9 μmol N�m−2�h−1).

Temporal variability of reef functions

Both C and N fluxes showed temporal variations;
however, significant differences in the magnitude and
directions occurred between coral- and algae-dominated
reef communities (Fig. 2, Table 1; detailed statistics in
Appendix S1: Table S4).
Daily-integrated GPP in coral-dominated communi-

ties remained stable throughout the year; however, CR
increased by >60% from winter to summer, resulting in
40% lower NCP (Table 1). GPP in algae-dominated
communities doubled from winter to summer, resulting
in significantly increased NCP that peaked at

>500 mmol C�m−2�d−1 in summer. In both community
types, variations in NCP were significantly correlated
with seawater temperature (Fig. 3a). NCP of coral-

Dark incubationsLight incubations

–8

0

0

80

160

2

40

–5

0
5

10

Ja
nu

ar
y

M
ar

ch
M

ay
Ju

ly
S

ep
te

m
be

r
N

ov
em

be
r

Ja
nu
ar
y
Ja
nu
ar
y
M
ar
ch
M
ay
Ju
ly
S
ep
te
m
be
r
N
ov
em
be
r
Ja
nu
ar
y

N
C

C

(
m

m
o

l C
a
C

O
m

·
h

)
D

O
C

(
m
m
o

l C
m

·h
)

D
IN

(
μ
m

o
l N

m
·
h

)

N

C
P

a
n
d

C
R

(
m
m
o
l C
m
·
h
)

–

20

0
20
40

Coral-dominated
Algae-dominated

(b)

(c)

(d)

(e) (f)

(g) (h)

(a)

Months

–

0.7

0
0.7

1.

4

FIG 2. Hourly biogeochemical fluxes during light (left panels)
and dark (right panels) incubations of coral- and algae-dominated
reef communities of the central Red Sea. Presented are all data
from bimonthly incubations of (a, b) community metabolism (di-
vided into net community production (NCP) and community res-
piration (CR), (c, d) net community calcification (NCC), (e, f) net
dissolved organic carbon (DOC) fluxes, and (g, h) net dissolved
inorganic nitrogen (DIN) fluxes. Dashed lines connect the
bimonthly means. Shaded areas connect the standard error around
the means. [Color figure can be viewed at wileyonlinelibrary.com]

Article e03226; page 6 FLORIAN ROTH ET AL. Ecology, Vol. 102, No. 2

www.wileyonlinelibrary.com

dominated communities exhibited a negative (r = −0.81,
P < 0.0001, n = 26), and NCP of algae-dominated com- munities exhibited a positive (r = 0.83, P < 0.0001, n = 26) relationship with increasing temperature. Hence, coral-dominated communities showed apparent negative Ea values for GPP and NCP, indicating them to be in the falling phase of the performance curve (i.e., past the opti- mum temperature), while apparent Ea for CR was posi- tive, indicating an opposite trend (Appendix S1: Fig. S3). In contrast, algae-dominated communities had positive Ea values for GPP, NCP, and, CR (Table 2), indicating that these communities remained within the rising phase of the performance curve throughout the study period (including the summer months; Appendix S1: Fig. S3). NCC in coral-dominated communities peaked in

spring, fall, and winter, with no differences between
these seasons, but dropped sharply by >60% in summer
(Table 1). Contrary, NCC in algae-dominated communi-
ties was consistently low throughout the year (Fig. 2c,d;
Table 1). NCC in coral-dominated communities corre-
lated negatively with increasing water temperatures
(r = −0.62, P = 0.0005, n = 26; Fig. 3b), negatively with
Ωarag (r = −0.36, P = 0.0070, n = 54) (Fig. 4a), and pos-
itively with NCP (r = 0.73, P < 0.0001, n = 54; Fig. 4b). NCC of algae-dominated communities did neither corre- late significantly with Ωarag, NCP, nor water tempera- ture. The corresponding apparent Ea values are presented in Table 2.

Net DOC fluxes per day in coral-dominated commu-
nities were lowest in winter and doubled in summer
(Table 1), owing to both increases in DOC releases dur-
ing dark and light incubations. While generally one
order of magnitude lower, net DOC fluxes integrated per
day remained stable in algae-dominated communities
throughout most of the year but increased towards sum-
mer (Table 1). Both community types showed a positive
correlation between net DOC fluxes and increasing tem-
perature (Fig. 3c), which was also reflected in positive
Ea values (Table 2, Appendix S1: Fig. S3).
DIN fluxes in both coral- and algae-dominated com-

munities were lowest in spring and twofold higher during
the rest of the year (Table 1). DIN fluctuations did not
correspond to changes in seawater temperature (Fig. 4d)
but showed a significant positive relationship (r = 0.57,
P = 0.0017, n = 26 for corals; r = 0.39, P = 0.0385,
n = 26 for algae) with increasing DIN concentrations of
the ambient seawater (Appendix S1: Fig. S2).

DISCUSSION

Phase shifts from corals to algae may drastically change
reef community functions

The carbon and carbonate pathways are well
described in the literature for undisturbed coral reef
communities, both in direction and magnitude

TABLE 1. Seasonal gross primary production (GPP), net community production (NCP), community respiration (CR), the ratio of
GPP and CR (GPP/CR), net community calcification (NCC), net dissolved organic carbon (DOC) fluxes, and net dissolved
inorganic nitrogen (DIN) fluxes of coral- and algae-dominated reef communities of the central Red Sea.

GPP (mmol
C⋅m⁻2⋅d⁻1)

NCP (mmol
C⋅m⁻2⋅d⁻1)

CR (mmol
C⋅m⁻2⋅d⁻1) GPP/CR

NCC (mmol
CaCO₃⋅m⁻2⋅d⁻1)

DOC (mmol
C⋅m⁻2⋅d⁻1)

DIN (mmol
N⋅m⁻2⋅d⁻1)

Annual mean
Coral 581 � 19 320 � 15 −261 � 17 2.4 � 0.1 131 � 10 14.3 � 1.8 1.6 � 0.2
Algae 946 � 51 443 � 21 −503 � 33 2.0 � 0.2 10 � 3 4.0 � 0.9 1.7 � 0.2
P <0.0001* <0.0001* <0.0001* <0.0001* <0.0001* <0.00001* 0.9351

Spring
Coral 638 � 21 353 � 12 −285 � 25 2.3 � 0.2 162 � 11 14.1 � 3.6 0.7 � 0.4
Algae 974 � 41 439 � 22 −535 � 27 1.8 � 0.0 7 � 6 1.6 � 1.6 1.3 � 0.4
∣t∣ <0.0001* 0.0833 <0.0001* 0.0134* <0.0001* 0.0246* 0.9573

Summer
Coral 575 � 45 231 � 24 −344 � 26 1.7 � 0.1 61 � 11 18.9 � 3.9 2.2 � 0.2
Algae 1,224 � 13 553 � 12 −671 � 9 1.8 � 0.0 13 � 3 7.1 � 1.9 1.9 � 0.4
∣t∣ <0.0001* <0.0001* <0.0001* 0.9310 0.0024 0.0432* 0.9979

Fall
Coral 447 � 44 292 � 25 −154 � 20 2.9 � 0.1 176 � 7 16.2 � 1.7 2.3 � 0.7
Algae 1,053 � 94 470 � 68 −583 � 43 1.8 � 0.1 15 � 11 2.3 � 3.0 2.3 � 0.7
∣t∣ <0.0001* 0.0017* <0.0001* <.00001* <0.0001* 0.0111* 1.0000

Winter
Coral 597 � 20 390 � 10 −207 � 15 2.9 � 0.1 147 � 10 9.1 � 3.0 1.6 � 0.3
Algae 586 � 30 324 � 17 −262 � 18 2.3 � 0.1 8 � 4 4.1 � 0.4 1.5 � 0.5
∣t∣ 1.0000 0.3351 0.6166 0.0005 <0.0001* 0.1388 1.0000

Notes: Values represent averages from all incubations during the respective season � SE. Significant differences between treat-
ments were assessed by linear mixed models (LMMs), and the differences between treatments and seasons by Tukey’s Honest Sig-
nificant Difference (HSD; Appendix S1: Table S4) test. Asterisks highlight significant P values (P or |t| < 0.05).

February 2021 FUNCTIONING OF CORAL REEF COMMUNITIES Article e03226; page 7

(Appendix S1: Table S5). Concomitant to previous find-
ings, coral-dominated communities in our experiments
displayed: (1) low net community production despite

high gross productivity, implying that biomass accumu-
lates slowly (Gattuso et al. 1998); (2) net DOC fluxes,
where the accumulation of exudates outpaces its con-
sumption (Nelson et al. 2013, Quinlan et al. 2018), rep-
resenting an organic C loss that further reduces C
available to support net benthic biomass accretion, and
promotes an efficient transfer of C via benthic-pelagic
coupling (Wild et al. 2004); and (3) high rates of net
community calcification, consistent with the accretion of
carbonate structures typical for tropical coral reefs (Gat-
tuso et al. 1998, 1999b, Atkinson and Falter 2003). Yet,
our seasonal comparative in situ approach revealed that
functions of reef ecosystems change with shifts from a
coral- to an algae-dominated benthic community.
Specifically, algae-dominated communities displayed a

higher organic C metabolism, along with residual net
calcification. Significantly higher GPP and >40% ele-
vated NCP in algae-dominated communities indicate a
greater potential for autotrophic biomass accumulation
per planar square meter of reef (Fong and Paul 2011,
Kelly et al. 2017), despite a lower GPP/CR ratio
(Table 1). The lower GPP/CR ratio in algae communi-
ties compared to coral communities seems counterintu-
itive considering the dominant organisms (algae as
autotrophs compared to mixotrophic corals); however,
algae-dominated reef structures host numerous

N
C

C
(
m

m
o
l C
a
C
O
m

·
d

)

26 28 30 32

(b)
0
80
160

240 ·d
)

26 28 30 32 32

(c)
0
20

40
N

C
P
(
m
m
o
l C
m
·
d
)
26 28 30 32

(a) ·d
)

26 28 30

(d)
0
2
4

Coral-dominated Algae-dominated

200

400

600

FIG 3. Relationship of temperature with (a) net community production (NCP), (b) net community calcification (NCC), (c) dis-
solved organic carbon (DOC) fluxes, and (d) dissolved inorganic nitrogen (DIN) fluxes in coral- and algae-dominated reef commu-
nities. Rates represent the net flux integrated over a 24-h day (assuming 12 h of light and 12 h of dark). Solid lines represent the
linear regressions; shaded areas in transparent colors represent the 95% confidence intervals. [Color figure can be viewed at wileyon
linelibrary.com]

TABLE 2. Apparent activation energies (Ea, eV) for coral- and algae-dominated reef communities as the slope of the Arrhenius
relationship between the natural logarithm of specific metabolic rates and the inverted temperature (1/kT).

GPP NCP CR NCC DOC DIN

Coral Algae Coral Algae Coral Algae Coral Algae Coral Algae Coral Algae

Ea (eV) −0.09 0.75 −0.58 0.55 0.49 0.93 −0.82 0.65 0.73 0.19 0.12 0.13
r2 0.03 0.71 0.54 0.59 0.26 0.66 0.30 0.10 0.28 0.01 0.02 0.02
P 0.3861 <0.0001 <0.0001 <0.0001 0.0065 <0.0001 0.0023 0.1073 0.0038 0.5619 0.5126 0.4439

Notes: GPP = gross primary production (mmol C�m−2�d−1); NCP = net community production (mmol C�m−2�d−1); CR = com-
munity respiration (mmol C�m−2�d−1); NCC = net community calcification = (mmol CaCO3�m−2�d−1); DOC = dissolved organic
carbon fluxes (mmol C�m−2�d−1); DIN = dissolved inorganic nitrogen fluxes (mmol N�m−2�d−1); r2 = square of correlation coeffi-
cient.

–5
0
5
10

–30 0 30 60

NCP (mmol C m ·h )

N
C
C
(
m
m
o
l C
a
C
O
m
·h
)

– 5

0
5
10

3 4 5 6

N
C
C
(
m
m
o
l C
a
C
O
m

·h
)

arag

(a) (b)

Coral-dominated Algae-dominated

FIG 4. Relationship between (a) net community calcifica-
tion (NCC) and aragonite saturation state (Ωarag), and (b) NCC
and net community production (NCP) in coral- and algae-dom-
inated reef communities. Closed circles indicate measurements
from light incubations, and open circles indicate dark incuba-
tions. Solid lines represent the linear regressions; shaded areas
in transparent colors represent the 95% confidence intervals.
[Color figure can be viewed at wileyonlinelibrary.com]

Article e03226; page 8 FLORIAN ROTH ET AL. Ecology, Vol. 102, No. 2

www.wileyonlinelibrary.com

www.wileyonlinelibrary.com

www.wileyonlinelibrary.com

heterotrophs that rely on the high algal biomass produc-
tion, fueling community-wide respiration by direct her-
bivory (Klumpp and McKinnon 1989, Russ 2003) and
indirect detritivory by invertebrates (Kramer et al.
2013). Likewise, sponges and other filter feeders are
commonly associated with degraded reef habitats (Abele
and Patton 1976), feeding on DOC or algal debris (Rix
et al. 2017, 2018). In addition, heterotrophic bacteria
within algal communities remineralize labile DOC
released by algae (Nelson et al. 2011, Haas et al. 2013).
Thereby, the consumption of algal-derived C (i.e., the C
retention within the system) can shorten the average
trophic path length, and reduce the average trophic level
of the second-order consumers (reviewed in Johnson
et al. 1995). In support of these trophic interactions, we
observed an apparent consumption of DOC in the dark
in algae-dominated communities, limiting the net DOC
flux integrated over 24 h (i.e., release in the light bal-
anced by consumption in the dark). Our results indicate
that algae-associated organisms readily remineralize
DOC, concurring with demonstrated DOC depletion in
algae-dominated shallow reefs elsewhere (Nelson et al.
2011, Haas et al. 2016).
Along with alterations of the organic C cycle, calcifi-

cation was reduced within algae-dominated communi-
ties. The slope of the relationship between NCP and
NCC is commonly used as an indicator of reef health,
indicative for the relative proportion of calcifying to
noncalcifying organisms in a benthic community (Alb-
right et al. 2013, Takeshita et al. 2016). We observed a
slope of 0.18 in coral- and 0.03 in algae-dominated com-
munities, and the slope averages 0.22 based on 52 reefs
around the world (Gattuso et al. 1999a). The significant
difference highlights the shift from calcifying corals to
noncalcifying organisms and a decoupling of the organic
carbon (production vs. respiration) and carbonate (calci-
fication vs. dissolution) cycles in algae-dominated com-
munities (McMahon et al. 2019).
Unraveling nitrogen pathways in coral reefs is crucial

to understand how high productivity is supported
despite low ambient nutrient concentrations (D’Elia and
Wiebe 1990, Szmant 2002, Atkinson and Falter 2003).
Both coral- and algae-dominated communities were net
sources of DIN over the study period with no significant
differences between community types. The flux rates
generally fell within the published range of in situ mea-
surements (−4 to 5 mmol N as NOx�m−2�d−1; reviewed
in Atkinson and Falter 2003). They are in stark contrast,
however, with the expectation that net autotrophic com-
munities would act as sinks for dissolved inorganic nutri-
ents, as generally measured in single organism
incubations (e.g., Den Haan et al. 2016). Although a the-
oretical N requirement of 34–83 mmol N�m−2�d−1 to
support NCP can be expected (stoichiometric calcula-
tions with NCP ranging from 230 to 550 mmol C�m−2-
�d−1; assuming a C/N ratio of 6.6; see Redfield 1958), the
effects of assimilation were likely masked by concurrent
community-wide processes that produce DIN (Gruber

et al. 2019). For example, cavities within Red Sea reefs
can be considerable sources of DIN (>20 mmol N�m−2-
�d−1) as sponges and other filter feeders utilize dissolved
organic matter (Richter et al. 2001). Also, microbial
communities can consume and transform organic N
compounds (Yahel et al. 2003, Moulton et al. 2016, Pfis-
ter and Altabet 2019), potentially increasing the commu-
nity-wide DIN release into the environment. Other
pathways, such as N2 fixation (Cardini et al. 2016) or
heterotrophic feeding on particulates (Ribes et al. 2003,
Houlbrèque and Ferrier-Pagès 2009) are additional N
sources that potentially limit/mask the N uptake from
DIN.
Overall, algae-dominated communities displayed a

higher potential of biomass accumulation or export (i.e.,
high NCP), associated with a higher total amount of C
available (i.e., high GPP) to the ecosystem. The high
NCP of algae communities facilitates rapid lateral and
vegetative overgrowth of bare substrates (Diaz-Pulido
and Garzón-Ferreira 2002, Roth et al. 2018). At the
same time, limited reef accretion (i.e., low NCC) within
algal habitats may compromise the topographic com-
plexity of phase-shifted reef communities (Wild et al.
2011), limit the recruitment of corals (Harrington et al.
2004, Roth et al. 2017, 2018), and increase reef erosion
(Adey 1978).

High temperatures during summer amplify functional
differences between coral- and algae-dominated

communities

Our data highlight that functions related to the carbon
and carbonate cycles of coral- and algae-dominated
communities are strongly but inversely affected by tem-
perature (summarized in Fig. 5), with implications for
their response to warming.
The results from activation energies show a higher

sensitivity to thermal stress of coral-dominated com-
pared to algae-dominated communities during summer
(Table 2, Appendix S1: Fig. S3). Particularly, apparent
activation energies for NCP and NCC of coral-domi-
nated communities (Ea = −0.58 and −0.82 eV; respec-
tively) were in the falling phase of the performance
curves, and thus, past the optimum temperature to peak
rates. Thereby, the community metabolism of corals is
pushed toward carbon losses because of high CR relative
to GPP in summer. Likewise, the activation energy of
CR was positive, as respiration continued to increase
with temperature through the seasonal thermal range.
Both the lower ratio of GPP to CR (GPP/CR) and
reduced NCP integrated over diel cycles indicate that
coral-dominated communities shifted toward a more
heterotrophic state with increasing temperature. It is
thus apparent that summer temperatures exceeded the
metabolic optima for coral-dominated communities,
which was previously suggested for individuals of Pocil-
lopora verrucosa (Sawall et al. 2015, Roik et al. 2016,
Anton et al. 2020) and Stylophora pistillata (Anton et al.

February 2021 FUNCTIONING OF CORAL REEF COMMUNITIES Article e03226; page 9

2020) in the central Red Sea. This contrasts with many
reef locations worldwide, where primary production
maxima are typically observed during the warmest
months of the year (e.g., Scheufen et al. 2017). At the
same time, coral-dominated communities displayed
enhanced rates of net DOC fluxes with warming (Ea =
0.73 eV), which can be attributed to an increased
release of cellular matter and/or mucoid exudates during
thermal stress in corals (Niggl et al. 2009, Scheufen et al.
2017). Although mucus released during higher tempera-
tures may help to protect corals against pathogens (Glasl
et al. 2016) or high UV radiation (Gleason and Welling-
ton 1993), it poses an increased loss of organic C that
can be used for community growth and/or export at con-
stant GPP rates (Fig. 5). Along with this trend, NCC

dropped by >50% from the annual mean during sum-
mer, with most of this decline realized as waters warmed
from 30 to 32°C. Overall, decreased NCC strongly corre-
lated with decreased NCP as temperatures increased (as
revealed by high negative apparent activation energies of
NCC, Ea = −0.82 eV), with a temperature threshold at
around 30.5°C (Appendix S1: Fig. S3). This thermal
threshold is near the reported thermal optimum of Pocil-
lopora verrucosa and Stylophora pistillata for gross pri-
mary production (29.9 and 31.9°C, respectively; Anton
et al. 2020) in the central Red Sea, indicating a strong
thermal sensitivity of coral-dominated communities
soon after the thermal optimum is exceeded. Accord-
ingly, in the temperature range above the optimum, the
rate of calcification decreased despite increased Ωarag in

FIG 5. In situ community metabolism of natural coral- and algae-dominated reef communities in the central Red Sea, Saudi
Arabia. Schematic was derived from all data available in the given lower (blue) and upper (red) temperature ranges. Organic carbon
pathways refer to photosynthesis, respiration, and dissolved organic carbon (DOC) fluxes, and the inorganic carbon pathway refers
to the formation and dissolution of calcium carbonate. The thicknesses of the bars scale with the actual flux measurements from
in situ incubations. [Color figure can be viewed at wileyonlinelibrary.com]

Article e03226; page 10 FLORIAN ROTH ET AL. Ecology, Vol. 102, No. 2

www.wileyonlinelibrary.com

summer (Fig. 4a; Silverman et al. 2007). Physiological
factors can also strongly affect the biomineralization
process. As calcification mainly depends on the photo-
synthetic efficiency of the endosymbionts within corals
(Gattuso et al. 1999a, Allemand et al. 2004), a lower
NCC may occur for thermally stressed corals, limiting
reef accretion and stabilization (Jokiel and Coles 1977,
De’ath et al. 2009).
In contrast, positive activation energies for algae-

dominated communities (Ea = 0.93, 0.75, and 0.55 eV
for CR, GPP, and NCP; respectively) indicate that these
communities benefit from higher temperatures and that
thermal optima were not reached in summer. In fact,
some macroalgae species (Halimeda tuna) from the cen-
tral Red Sea have a reported thermal optimum (31.7°C)
for gross primary production (Anton et al. 2020) that is
close to the maximum summer temperature recorded
during our incubations (32.5°C). The increases in GPP
and CR along a thermal gradient in algae-dominated
communities highlight a higher turnover of organic C.
However, increases in C fixation outweighed increases in
respiratory C consumption, resulting in higher NCP in
the summer. In contrast to previous reports (e.g., Barron
et al. 2014), DOC fluxes in algae-dominated communi-
ties showed only a weak temperature dependence. How-
ever, although the overall net DOC fluxes remained
relatively stable, differences in net production during the
light and net consumption in the dark amplified with
temperature (Fig. 3), limiting losses of organic C
through this process. As a result, more organic C was
retained within algae-dominated communities in sum-
mer, supporting biomass accumulation and export in the
community (Fig. 5).

Implications for reef ecosystem functioning under global
change

Theoretical studies have provided a sound under-
standing of the relationship between biodiversity loss
and ecosystem functions in tropical coral reefs (reviewed
in Hughes et al. 2017). However, considerable knowledge
gaps remain, in particular, on how metabolic and bio-
geochemical processes differ between coral- and algae-
dominated communities, and how these respond to sea-
sonal fluctuations in environmental conditions. As algal
cover is expected to increase in coral reefs, our long-term
in situ experiments reveal how these novel communities
in general, and how thermal stress in particular, may
alter pivotal ecological functions of future reefs. Our
data show that fundamental metabolic and biogeochem-
ical characteristics of coral-dominated communities are
disturbed by shifts from coral to algal dominance and
may, thereby, compromise the future stability and resili-
ence of coral reef biota. These responses may be further
compounded by differential thermal responses between
coral and algae species (e.g., Anton et al. 2020).
The sensitivity of corals and their symbionts to rising

temperatures has been documented extensively (Hoegh-

Guldberg 1999). Thermal anomalies exceeding 1–2°C
above the mean summer maximum temperature can
compromise the symbiosis (e.g., Weeks et al. 2008), lead-
ing to coral bleaching and reduced coral survival (Baird
and Marshall 2002, Baker et al. 2008). However, our
study did not record temperatures exceeding the local
mean summer maxima reported for the region (see
Fig. 1c; Chaidez et al. 2017) and, likewise, no apparent
signs of coral bleaching were observed. Nevertheless,
growth of Pocillopora verrucosa and Stylophora pistillata
(Anton et al. 2020) and calcification of Pocillopora verru-
cosa (Roik et al. 2018) are already reduced under current
summer conditions in the Red Sea, as also highlighted
by the present study. For the warmer part of the year,
algae-dominated communities have, thus, a metabolic
advantage over coral-dominated communities because
they maintain high NCP. Importantly, the consequences
of global warming may manifest not only in terms of
higher-than-normal temperatures but also in a longer-
than-normal duration of the seasonal peak temperatures
(Fitt et al. 2001). Under such conditions, if coral mortal-
ity events occur, algae may quickly spread and shift coral
ecosystems more rapidly towards systems that are domi-
nated by algae (McManus et al. 2019), as already
reported on some reefs in the southern Red Sea follow-
ing the coral bleaching event in 2015 (Anton et al. 2020).
The frequency and intensity of climate-driven stress

events on coral reefs will inevitably aggravate in the near
future. Our results suggest that the anticipated increase
in the spatial footprint of algae-dominated communities
would exacerbate the magnitude of the functional
changes described here. Ocean warming likely enhances
the competitive advantage of algae- over coral-domi-
nated communities (Anton et al. 2020), thus promoting
a positive feedback loop of reef degradation. Similar
effects of warming are likely to be operational in other
temperature-sensitive and calcifying communities. As
these organisms are central to the formation of reef
ecosystems, critical changes in the biodiversity and func-
tioning may be witnessed. Thus, appropriate manage-
ment practices designed to limit the proliferation of
algae are needed for maintaining reefs dominated by
corals and the important ecosystem services they sup-
port.

ACKNOWLEDGMENTS

We are grateful to the personnel from the King Abdullah
University of Science and Technology (KAUST) Coastal and
Marine Resources Core (CMOR) Laboratory for logistical sup-
port. The authors would also like to acknowledge Rodrigo Vil-
lalobos and João Cúrdia, who helped during fieldwork.
Figure 5 was produced by Xavier Pita, scientific illustrator at
KAUST. We would like to thank the two anonymous reviewers
and the editor for their helpful suggestions and comments. The
research was supported by KAUST baseline funding to BHJ
and by grant Wi 2677/9-1 from the German Research Founda-
tion (DFG) to CW. Author contributions: FR, CW, and SC
conceptualized and designed research. FR, LS, MLC, and VS
performed research. FR, NR, VS, AA, LS, BK, and MLC

February 2021 FUNCTIONING OF CORAL REEF COMMUNITIES Article e03226; page 11

analyzed data. CMD, XAGM, CRV, and BHJ contributed to
research materials, logistics and to interpreting data. FR wrote
original draft of the manuscript with support by CW. All
authors read and approved the final manuscript.

LITERATURE CITED

Abele, L. G., and W. K. Patton. 1976. The size of coral heads
and the community biology of associated decapod crus-
taceans. Journal of Biogeography 3:35.

Adey, W. H. 1978. Coral reef morphogensis: a multidimensional
model. Science 202(4370): 831–837.

Albright, R., C. Langdon, and K. R. N. Anthony. 2013.
Dynamics of seawater carbonate chemistry, production, and
calcification of a coral reef flat, Central Great Barrier Reef.
Biogeosciences Discussions 10:7641–7676.

Allemand, D., C. Ferrier-Pagès, P. Furla, F. Houlbrèque, S.
Puverel, S. Reynaud, É. Tambutté, S. Tambutté, and D. Zoc-
cola. 2004. Biomineralisation in reef-building corals: from
molecular mechanisms to environmental control. Comptes
Rendus Palevol 3:453–467.

Anthony, K. R. N., D. I. Kline, G. Diaz-Pulido, S. Dove, and O.
Hoegh-Guldberg. 2008. Ocean acidification causes bleaching
and productivity loss in coral reef builders. Proceedings of the
National Academy of Sciences of the United States of Amer-
ica 105:17442–17446.

Anton, A., J. L. Randle, F. C. Garcia, S. Rossbach, J. I. Ellis, M.
Weinzierl, and C. M. Duarte. 2020. Differential thermal toler-
ance between algae and corals may trigger the proliferation of
algae in coral reefs. Global Change Biology 26:4316–4327.

Ateweberhan, M., J. H. Bruggemann, and A. M. Breeman.
2006. Effects of extreme seasonality on community structure
and functional group dynamics of coral reef algae in the
southern Red Sea (Eritrea). Coral Reefs 25:391–406.

Atkinson, M. J., and J. L. Falter. 2003. Coral reefs. in K. Black
and G. Shimmieldeditors. Pages 40–64 Biogeochemistry of
Marine Systems, Boca Raton, Florida: CRC Press.

Baird, A. H., and P. A. Marshall. 2002. Mortality, growth and
reproduction in scleractinian corals following bleaching on
the Great Barrier Reef. Marine Ecology Progress Series
237:133–141.

Baird, D., S. Brown, L. Lagadic, M. Liess, L. Maltby, M. Mor-
eira-Santos, R. Schulz, and G. Scott. 2007. In situ–based
effects measures: determining the ecological relevance of mea-
sured responses. Integrated Environmental Assessment and
Management 3:259–267.

Baker, A. C., P. W. Glynn, and B. Riegl. 2008. Climate change
and coral reef bleaching: an ecological assessment of long-
term impacts, recovery trends and future outlook. Estuarine,
Coastal and Shelf Science 80:435–471.

Balvanera, P., A. B. Pfisterer, N. Buchmann, J. S. He, T. Naka-
shizuka, D. Raffaelli, and B. Schmid. 2006. Quantifying the
evidence for biodiversity effects on ecosystem functioning
and services. Ecology Letters 9:1146–1156.

Barron, C., E. T. Apostolaki, and C. M. Duarte. 2014. Dis-
solved organic carbon fluxes by seagrass meadows and
macroalgal beds. Frontiers in Marine Science 1:42.

Bellwood, D. R., T. P. Hughes, C. Folke, and M. Nyström.
2004. Confronting the coral reef crisis. Nature 429:827–833.

Bellworthy, J., and M. Fine. 2018. The Red Sea Simulator: A
high-precision climate change mesocosm with automated
monitoring for the long-term study of coral reef organisms.
Limnology and Oceanography: Methods 16:367–375.

Bender, D., G. Diaz-Pulido, and S. Dove. 2014. Warming and
acidification promote cyanobacterial dominance in turf algal
assemblages. Marine Ecology Progress Series 517:271–284.

Bernstein, W. N., K. A. Hughen, C. Langdon, D. C. McCorkle,
and S. J. Lentz. 2016. Environmental controls on daytime net
community calcification on a Red Sea reef flat. Coral Reefs
35:697–711.

Brondizio, E., J. Settele, S. Diaz, and T. Ngo. 2019. Global
assessment report on biodiversity and ecosystem services of
the Intergovernmental Science—Policy Platform on Biodiver-
sity and Ecosystem Services. IPBES Secretariat, Bonn, Ger-
many.

Bythell, J. C., Z. M. Hillis-Starr, and C. S. Rogers. 2000. Local
variability but landscape stability in coral reef communities
following repeated hurricane impacts. Marine Ecology Pro-
gress Series 204:93–100.

Canavero, A., M. Arim, F. Pérez, F. M. Jaksic, and P. A. Mar-
quet. 2018. A metabolic view of amphibian local community
structure: the role of activation energy. Ecography
41:388–400.

Cardini, U., V. N. Bednarz, N. van Hoytema, A. Rovere, M. S.
Naumann, M. M. D. Al-Rshaidat, and C. Wild. 2016. Budget
of primary production and dinitrogen fixation in a highly sea-
sonal Red Sea coral reef. Ecosystems 19:771–785.

Chaidez, V., D. Dreano, S. Agusti, C. M. Duarte, and I. Hoteit.
2017. Decadal trends in Red Sea maximum surface tempera-
ture. Scientific Reports 7:1–8.

D’Elia, C. F., and W. J. Wiebe. 1990. Biogeochemical nutrient
cycles in coral reef ecosystems. Pages 49–74 in Z. Dubinsky,
editor. Ecosystems of the world 25: coral reefs. Elsevier, New
York, New York, USA.

Damgaard, C. 2019. A critique of the space-for-time substitu-
tion practice in community ecology. Trends in Ecology and
Evolution 34:416–421.

de Jongh, F., and F. C. Van Duyl. 2004. The impact of suspen-
sion feeders in cryptic habitats on the trophodynamics of
coral reefs. Thesis. http://www.science.uva.nl/ZMA/Inverteb
rates/Coel/scirep/Cryptic_Reef_Habitats

De’ath, G., J. M. Lough, and K. E. Fabricius. 2009. Declining
coral calcification on the Great Barrier Reef. Science
323:116–119.

Den Haan, J., et al. 2016. Nitrogen and phosphorus uptake
rates of different species from a coral reef community after a
nutrient pulse. Scientific Reports 6:1–13.

Diaz-Pulido, G., and J. Garzón-Ferreira. 2002. Seasonality in
algal assemblages on upwelling-influenced coral reefs in the
Colombian Caribbean. Botanica Marina 45:284–292.

Done, T. J. 1992. Phase shifts in coral reef communities and
their ecological significance. in V. Jaccarini and E. Martens
editors. Pages 121–132 Book: The ecology of mangrove and
related ecosystems, Dordrecht, Netherlands: Springer Inter-
national Publishing.

Done, T. J., P. K. Dayton, A. E. Dayton, and R. Steger. 1991.
Regional and local variability in recovery of shallow coral
communities: Moorea, French Polynesia and central Great
Barrier Reef. Coral Reefs 9:183–192.

Edmunds, P. J. 2002. Long-term dynamics of coral reefs in St.
John, US Virgin Islands. Coral Reefs 21:357–367.

Edmunds, P. J., S. S. Doo, and R. C. Carpenter. 2020. Year-long
effects of high pCO2 on the community structure of a tropical
fore reef assembled in outdoor flumes. ICES Journal of Mar-
ine Science 77:1055–1065.

Fitt, W., B. Brown, M. Warner, and R. Dunne. 2001. Coral
bleaching: interpretation of thermal tolerance limits and ther-
mal thresholds in tropical corals. Coral Reefs 20:51–65.

Follstad Shah, J. J., et al. 2017. Global synthesis of the tempera-
ture sensitivity of leaf litter breakdown in streams and rivers.
Global Change Biology 23:3064–3075.

Fong, P., and V. J. Paul. 2011. Coral reef algae. in Z. Dubinsky
and N Stambler editors. Pages 241–272 Book: Coral reefs: an

Article e03226; page 12 FLORIAN ROTH ET AL. Ecology, Vol. 102, No. 2

http://www.science.uva.nl/ZMA/Invertebrates/Coel/scirep/Cryptic_Reef_Habitats

http://www.science.uva.nl/ZMA/Invertebrates/Coel/scirep/Cryptic_Reef_Habitats

ecosystem in transition, Dordrecht, Netherlands: . Springer
International Publishing.

Frölicher, T. L., E. M. Fischer, and N. Gruber. 2018. Marine
heatwaves under global warming. Nature 560:360–364.

Garcı́a, F. C., E. Bestion, R. Warfield, and G. Yvon-Durocher.
2018. Changes in temperature alter the relationship between
biodiversity and ecosystem functioning. Proceedings of the
National Academy of Sciences of the United States of Amer-
ica 115:10989–10994.

Gattuso, J. P., D. Allemand, and M. Frankignoulle. 1999a. Pho-
tosynthesis and calcification at cellular, organismal and com-
munity levels in coral reefs: a review on interactions and
control by carbonate chemistry. American Zoologist
39:160–183.

Gattuso, J. P., M. Frankignoulle, and S. V. Smith. 1999b. Mea-
surement of community metabolism and significance in the
coral reef CO2 source-sink debate. Proceedings of the
National Academy of Sciences of the United States of Amer-
ica 96:13017–13022.

Gattuso, J. P., M. Frankignoulle, and R. Wollast. 1998. Carbon
and carbonate metabolism in coastal aquatic ecosystems.
Annual Review of Ecology and Systematics 29:405–434.

Glasl, B., G. J. Herndl, and P. R. Frade. 2016. The microbiome
of coral surface mucus has a key role in mediating holobiont
health and survival upon disturbance. ISME Journal
10:2280–2292.

Gleason, D. F., and G. M. Wellington. 1993. Ultraviolet radia-
tion and coral bleaching. Nature 365:836–838.

Graham, N. A. J., S. Jennings, M. A. MacNeil, D. Mouillot,
and S. K. Wilson. 2015. Predicting climate-driven regime
shifts versus rebound potential in coral reefs. Nature
518:1–17.

Gruber, R. K., R. J. Lowe, and J. L. Falter. 2019. Tidal and sea-
sonal forcing of dissolved nutrient fluxes in reef communities.
Biogeosciences 16 1921–1935. https://doi.org/10.5194/bg-16-
1921-2019.

Haas, A. F., et al. 2016. Global microbialization of coral reefs.
Nature Microbiology 1: 16042.

Haas, A. F., C. E. Nelson, F. Rohwer, L. Wegley-Kelly, S. D.
Quistad, C. A. Carlson, J. J. Leichter, M. Hatay, and J. E.
Smith. 2013. Influence of coral and algal exudates on micro-
bially mediated reef metabolism. PeerJ 1:e108.

Harrington, L., K. Fabricius, G. D’eath, A. Negri, L. I. H.
Arrington, K. A. F. Abricius, and G. L. D. E. Ath. 2004.
Recognition and selection of settlement substrata determine
post-settlement survival in corals. Ecology 85:3428–3437.

Hempson, T. N., N. A. J. Graham, M. A. Macneil, A. S. Hoey,
and S. K. Wilson. 2018. Ecosystem regime shifts disrupt
trophic structure. Ecological Applications 28:191–200.

Hoegh-Guldberg, O. 1999. Climate change, coral bleaching and
the future of the world’s coral reefs. Marine and Freshwater
Research 50:839–866.

Hooper, D. U., E. C. Adair, B. J. Cardinale, J. E. K. Byrnes, B.
A. Hungate, K. L. Matulich, A. Gonzalez, J. E. Duffy, L.
Gamfeldt, and M. I. Connor. 2012. A global synthesis reveals
biodiversity loss as a major driver of ecosystem change. Nat-
ure 486:105–108.

Houlbrèque, F., and C. Ferrier-Pagès. 2009. Heterotrophy in
tropical scleractinian corals. Biological Reviews 84:1–17.

Hughes, T. P., et al. 2017. Coral reefs in the Anthropocene. Nat-
ure 546:82–90.

Hughes, T. P., et al. 2019. Ecological memory modifies the
cumulative impact of recurrent climate extremes. Nature Cli-
mate Change 9:40–43.

Hughes, T. P., M. J. Rodrigues, D. R. Bellwood, D. Ceccarelli,
O. Hoegh-Guldberg, L. McCook, N. Moltschaniwskyj, M. S.
Pratchett, R. S. Steneck, and B. Willis. 2007. Phase shifts,

herbivory, and the resilience of coral reefs to climate change.
Current Biology 17:360–365.

Jessen, C., J. F. Villa Lizcano, T. Bayer, C. Roder, M. Aranda,
C. Wild, and C. R. Voolstra. 2013. In-situ effects of eutrophi-
cation and overfishing on physiology and bacterial diversity
of the red sea coral Acropora hemprichii. PLoS One 8:e62091.

Johnson, C., D. Klumpp, J. Field, and R. Bradbury. 1995. Car-
bon flux on coral reefs: effects of large shifts in community
structure. Marine Ecology Progress Series 126:123–143.

Jokiel, P. L., and S. L. Coles. 1977. Effects of temperature on
the mortality and growth of Hawaiian reef corals. Marine
Biology 43:201–208.

Kelly, E. L. A., Y. Eynaud, I. D. Williams, R. T. Sparks, M. L. Dai-
ler, S. A. Sandin, and J. E. Smith. 2017. A budget of algal pro-
duction and consumption by herbivorous fish in an herbivore
fisheries management area, Maui. Hawaii. Ecosphere 8:e01899.

Klumpp, W., and A. McKinnon. 1989. Community structure,
biomass and productivity of epilithic algal communities on
the Great Barrier Reef: dynamics at different spatial scales.
Marine Ecology Progress Series 86:77–89.

Koch, M., G. Bowes, C. Ross, and X. H. Zhang. 2013. Climate
change and ocean acidification effects on seagrasses and mar-
ine macroalgae. Global Change Biology 19:103–132.

Kramer, M. J., O. Bellwood, and D. R. Bellwood. 2013. The
trophic importance of algal turfs for coral reef fishes: the
crustacean link. Coral Reefs 32:575–583.

Langdon, C., W. S. Broecker, D. E. Hammond, E. Glenn, K.
Fitzsimmons, S. G. Nelson, T.-H. Peng, I. Hajdas, and G.
Bonani. 2003. Effect of elevated CO2 on the community
metabolism of an experimental coral reef. Global Biogeo-
chemical Cycles 17:1011.

Lavigne, H., and J. P. Gattuso. 2013. Package “seacarb”: seawa-
ter carbonate chemistry with R, v. 2.4. 8. http://www2.uaem.
mx/r-mirror/web/packages/seacarb/seacarb

Legendre, P., and L. Legendre. 2012. Numerical ecology. Third
edition. Elsevier, Amsterdam, The Netherlands.

Lirman, D., and P. Biber. 2000. Seasonal dynamics of macroal-
gal communities of the northern Florida reef tract. Botanica
Marina 43:305–314.

Lough, J. M., K. D. Anderson, and T. P. Hughes. 2018. Increas-
ing thermal stress for tropical coral reefs: 1871–2017. Scien-
tific Reports 8:6079.

Maldonado, M., M. Ribes, and F. C. van Duyl. 2012. Nutrient
fluxes through sponges. Biology, budgets, and ecological
implications. Advances in Marine Biology 62:113–182.

McMahon, A., I. R. Santos, T. Cyronak, and B. D. Eyre. 2013.
Hysteresis between coral reef calcification and the seawater
aragonite saturation state. Geophysical Research Letters
40:4675–4679.

McMahon, A., I. R. Santos, K. G. Schulz, A. Scott, J. Silver-
man, K. L. Davis, and D. T. Maher. 2019. Coral reef calcifica-
tion and production after the 2016 bleaching event at Lizard
Island, Great Barrier reef. Journal of Geophysical Research
Oceans 124:4003–4016.

McManus, L. C., V. V. Vasconcelos, S. A. Levin, D. M. Thomp-
son, J. A. Kleypas, F. S. Castruccio, E. N. Curchitser, and J.
R. Watson. 2019. Extreme temperature events will drive coral
decline in the Coral Triangle. Global Change Biology
26:2120–2133.

Middleton, B., and J. Grace. 2004. Biodiversity and ecosystem
functioning: synthesis and perspectives. Restoration Ecology
12:611–612.

Moberg, F. F., and C. Folke. 1999. Ecological goods and ser-
vices of coral reef ecosystems. Ecological Economics
29:215–233.

Morán, X. A. G., J. M. Gasol, M. C. Pernice, J. Mangot, R.
Massana, E. Lara, D. Vaqué, and C. M. Duarte. 2017.

February 2021 FUNCTIONING OF CORAL REEF COMMUNITIES Article e03226; page 13

https://doi.org/10.5194/bg-16-1921-2019

https://doi.org/10.5194/bg-16-1921-2019

http://www2.uaem.mx/r-mirror/web/packages/seacarb/seacarb

http://www2.uaem.mx/r-mirror/web/packages/seacarb/seacarb

Temperature regulation of marine heterotrophic prokaryotes
increases latitudinally as a breach between bottom-up and
top-down controls. Global Change Biology 23:3956–3964.

Moulton, O. M., M. A. Altabet, J. M. Beman, L. A. Deegan, J.
Lloret, M. K. Lyons, J. A. Nelson, and C. A. Pfister. 2016.
Microbial associations with macrobiota in coastal ecosys-
tems: patterns and implications for nitrogen cycling. Frontiers
in Ecology and the Environment 14:200–208.

Naeem, S., J. E. Duffy, and E. Zavaleta. 2012. The functions of
biological diversity in an age of extinction. Science
336:1401–1406.

Nelson, C. E., A. L. Alldredge, E. A. McCliment, L. A.
Amaral-Zettler, and C. A. Carlson. 2011. Depleted dissolved
organic carbon and distinct bacterial communities in the
water column of a rapid-flushing coral reef ecosystem. ISME
Journal 5:1374–1387.

Nelson, C. E., S. J. Goldberg, L. Wegley Kelly, A. F. Haas, J. E.
Smith, F. Rohwer, and C. A. Carlson. 2013. Coral and
macroalgal exudates vary in neutral sugar composition and
differentially enrich reef bacterioplankton lineages. ISME
Journal 7:962–979.

Niggl, W., M. Glas, C. Laforsch, Mayr, C., and Wild, C. 2009.
First evidence of coral bleaching stimulating organic matter
release by reef corals. Pages 905–910inB. Riegl, editor. Pro-
ceedings of the 11th International Coral Reef Symposium.
https://www.researchgate.net/profile/Martin_Gutbrod2/publi
cation/257233311_First_evidence_of_coral_bleaching_stimu
lating_organic_matter_release_by_reef_corals/links/56938980
08aec14fa55e9224/First-evidence-of-coral-bleaching-stimula
ting-organic-matter-release-by-reef-corals .

Oliver, E. C. J., et al. 2018. Longer and more frequent marine
heatwaves over the past century. Nature Communications
9:1–12.

Padfield, D., C. Lowe, A. Buckling, R. Ffrench-Constant, S.
Jennings, F. Shelley, J. S. Ólafsson, and G. Yvon-Durocher.
2017. Metabolic compensation constrains the temperature
dependence of gross primary production. Ecology Letters
20:1250–1260.

Pfister, C. A., and M. A. Altabet. 2019. Enhanced microbial
nitrogen transformations in association with macrobiota
from the rocky intertidal. Biogeosciences 16:193–206.

Pierrot, D., E. Lewis, and D. W. R. Wallace. 2006. MS Excel
program developed for CO2 system calculations. Oak Ridge,
Tennessee, USA. https://www.ncei.noaa.gov/access/ocean-ca
rbon-data-system/oceans/CO2SYS/co2rprt.html.

Piggot, A. M., B. W. Fouke, M. Sivaguru, R. A. Sanford, and
H. R. Gaskins. 2009. Change in zooxanthellae and mucocyte
tissue density as an adaptive response to environmental stress
by the coral, Montastraea annularis. Marine Biology
156:2379–2389.

Quinlan, Z. A., K. Remple, M. D. Fox, N. J. Silbiger, T. A. Oli-
ver, H. M. Putnam, L. Wegley Kelly, C. A. Carlson, M. J.
Donahue, and C. E. Nelson. 2018. Fluorescent organic exu-
dates of corals and algae in tropical reefs are compositionally
distinct and increase with nutrient enrichment. Limnology
and Oceanography Letters 3:331–340.

Redfield, A. C. 1958. The biological control of chemical factors
in the environment. American Scientist 46:230A–A221.

Reynaud, S., N. Leclercq, S. Romaine-Lioud, C. Ferrier-Pagès,
J. Jaubert, and J. P. Gattuso. 2003. Interacting effects of CO2
partial pressure and temperature on photosynthesis and calci-
fication in a scleractinian coral. Global Change Biology
9:1660–1668.

Ribes, M., R. Coma, M. Atkinson, and R. Kinzie. 2003. Parti-
cle removal by coral reef communities: picoplankton is a
major source of nitrogen. Marine Ecology Progress Series
257:13–23.

Richter, C., and M. Wunsch. 1999. Cavity-dwelling suspension
feeders in coral reefs—A new link in reef trophodynamics.
Marine Ecology Progress Series 188:105–116.

Richter, C., M. Wunsch, M. Rasheed, I. Kötter, and M. I. Bad-
ran. 2001. Endoscopic exploration of Red Sea coral reefs
reveals dense populations of cavity-dwelling sponges. Nature
413:726–730.

Rix, L., V. N. Bednarz, U. Cardini, N. Van Hoytema, F. A. Al-
Horani, C. Wild, and M. S. Naumann. 2015. Seasonality in
dinitrogen fixation and primary productivity by coral reef
framework substrates from the northern Red Sea. Marine
Ecology Progress Series 533:79–92.

Rix, L., J. M. De Goeij, D. Van Oevelen, U. Struck, F. A. Al-
Horani, C. Wild, and M. S. Naumann. 2018. Reef sponges
facilitate the transfer of coral-derived organic matter to their
associated fauna via the sponge loop. Marine Ecology Pro-
gress Series 589:85–96.

Rix, L., et al. 2017. Differential recycling of coral and algal dis-
solved organic matter via the sponge loop. Functional Ecol-
ogy 31:778–789.

Roik, A., C. Roder, T. Röthig, and C. R. Voolstra. 2016. Spatial
and seasonal reef calcification in corals and calcareous crusts
in the central Red Sea. Coral Reefs 35:681–693.

Roik, A., T. Röthig, C. Pogoreutz, V. Saderne, and C. R. Vool-
stra. 2018. Coral reef carbonate budgets and ecological dri-
vers in the central Red Sea—A naturally high temperature
and high total alkalinity environment. Biogeosciences
15:6277–6296.

Roth, F., F. Saalmann, T. Thomson, D. J. Coker, R. Villalobos,
B. H. Jones, C. Wild, and S. Carvalho. 2018. Coral reef degra-
dation affects the potential for reef recovery after disturbance.
Marine Environmental Research 142:48–58.

Roth, F., I. Stuhldreier, C. Sánchez-Noguera, S. Carvalho, and
C. Wild. 2017. Simulated overfishing and natural eutrophica-
tion promote the relative success of a non-indigenous ascid-
ian in coral reefs at the pacific coast of Costa Rica. Aquatic
Invasions 12:435–446.

Roth, F., C. Wild, S. Carvalho, N. Rädecker, C. R. Voolstra, B.
Kürten, H. Anlauf, Y. C. El-Khaled, R. Carolan, and B. H.
Jones. 2019. An in situ approach for measuring biogeochemi-
cal fluxes in structurally complex benthic communities. Meth-
ods in Ecology and Evolution 10:712–725.

Russ, G. R. 2003. Grazer biomass correlates more strongly with
production than with biomass of algal turfs on a coral reef.
Coral Reefs 22:63–67.

Savva, I., S. Bennett, G. Roca, G. Jordà, and N. Marbà. 2018.
Thermal tolerance of Mediterranean marine macrophytes:
Vulnerability to global warming. Ecology and Evolution
8:12032–12043.

Sawall, Y., A. Al-Sofyani, S. Hohn, E. Banguera-Hinestroza, C.
R. Voolstra, and M. Wahl. 2015. Extensive phenotypic plas-
ticity of a Red Sea coral over a strong latitudinal temperature
gradient suggests limited acclimatization potential to warm-
ing. Scientific Reports 5:8940.

Scheufen, T., W. E. Krämer, R. Iglesias-Prieto, and S. Enrı́quez.
2017. Seasonal variation modulates coral sensibility to heat-
stress and explains annual changes in coral productivity. Sci-
entific Reports 7:4937.

Sibly, R., J. Brown, and A. Kodric-Brown. 2012. Metabolic
ecology: a scaling approach. Wiley-Blackwell: Oxford, UK
ISBN: 978-1-119-96851-1.

Silverman, J., B. Lazar, and J. Erez. 2007. Effect of aragonite
saturation, temperature, and nutrients on the community cal-
cification rate of a coral reef. Journal of Geophysical
Research: Oceans 112:1–14.

Stachowicz, J. J., J. F. Bruno, and J. E. Duffy. 2007. Understand-
ing the effects of marine biodiversity on communities and

Article e03226; page 14 FLORIAN ROTH ET AL. Ecology, Vol. 102, No. 2

https://www.researchgate.net/profile/Martin_Gutbrod2/publication/257233311_First_evidence_of_coral_bleaching_stimulating_organic_matter_release_by_reef_corals/links/5693898008aec14fa55e9224/First-evidence-of-coral-bleaching-stimulating-organic-matter-release-by-reef-corals

https://www.researchgate.net/profile/Martin_Gutbrod2/publication/257233311_First_evidence_of_coral_bleaching_stimulating_organic_matter_release_by_reef_corals/links/5693898008aec14fa55e9224/First-evidence-of-coral-bleaching-stimulating-organic-matter-release-by-reef-corals

https://www.researchgate.net/profile/Martin_Gutbrod2/publication/257233311_First_evidence_of_coral_bleaching_stimulating_organic_matter_release_by_reef_corals/links/5693898008aec14fa55e9224/First-evidence-of-coral-bleaching-stimulating-organic-matter-release-by-reef-corals

https://www.researchgate.net/profile/Martin_Gutbrod2/publication/257233311_First_evidence_of_coral_bleaching_stimulating_organic_matter_release_by_reef_corals/links/5693898008aec14fa55e9224/First-evidence-of-coral-bleaching-stimulating-organic-matter-release-by-reef-corals

https://www.researchgate.net/profile/Martin_Gutbrod2/publication/257233311_First_evidence_of_coral_bleaching_stimulating_organic_matter_release_by_reef_corals/links/5693898008aec14fa55e9224/First-evidence-of-coral-bleaching-stimulating-organic-matter-release-by-reef-corals

https://www.ncei.noaa.gov/access/ocean-carbon-data-system/oceans/CO2SYS/co2rprt.html

https://www.ncei.noaa.gov/access/ocean-carbon-data-system/oceans/CO2SYS/co2rprt.html

ecosystems. Annual Review of Ecology, Evolution, and Sys-
tematics 38:739–766.

Stillman, J. H. 2019. Heat waves, the new normal: summertime
temperature extremes will impact animals, ecosystems, and
human communities. Physiology 34:86–100.

Szmant, A. M. 2002. Nutrient enrichment on coral reefs: is
it a major cause of coral reef decline? Estuaries 25:
743–766.

Takeshita, Y., W. Mcgillis, E. M. Briggs, A. L. Carter, E. M.
Donham, T. R. Martz, N. N. Price, and J. E. Smith. 2016.
Assessment of net community production and calcification of
a coral reef using a boundary layer approach. Journal of Geo-
physical Research: Oceans 121:5655–5671.

Tkachenko, K. S., B. J. Wu, L. S. Fang, and T. Y. Fan. 2007.
Dynamics of a coral reef community after mass mortality of
branching Acropora corals and an outbreak of anemones.
Marine Biology 151:185–194.

Van Heuven, S. M. A. C., A. E. Webb, D. M. de Bakker, E.
Meesters, F. C. Van Duyl, G.-J.-J. Reichart, and L. J. De
Nooijer. 2018. In-situ incubation of a coral patch for commu-
nity-scale assessment of metabolic and chemical processes on
a reef slope. PeerJ 2018:e5966.

Weeks, S. J., K. R. N. Anthony, A. Bakun, G. C. Feldman, and
O.-H. Guldberg. 2008. Improved predictions of coral bleach-
ing using seasonal baselines and higher spatial resolution.
Limnology and Oceanography 53:1369–1375.

Wild, C., et al. 2011. Climate change impedes scleractinian cor-
als as primary reef ecosystem engineers. Marine and Freshwa-
ter Research 62:205–215.

Wild, C., M. Huettel, A. Klueter, S. G. Kremb, M. Y. M.
Rasheed, and B. B. Jørgensen. 2004. Coral mucus functions
as an energy carrier and particle trap in the reef ecosystem.
Nature 428:66–70.

Wolf-Gladrow, D. A., R. E. Zeebe, C. Klaas, A. Körtzinger,
and A. G. Dickson. 2007. Total alkalinity: The explicit con-
servative expression and its application to biogeochemical
processes. Marine Chemistry 106:287–300.

Yahel, G., J. H. Sharp, D. Marie, C. Häse, and A. Genin. 2003.
In situ feeding and element removal in the symbiont-bearing
sponge Theonella swinhoei: Bulk DOC is the major source for
carbon. Limnology and Oceanography 48:141–149.

Zeebe, R. E., and D. A. Wolf-Gladrow. 2001. CO2 in seawater :
equilibrium, kinetics, isotopes. First editionAmsterdam,
Netherlands: Elsevier Science. ISBN: 9780080929903

SUPPORTING INFORMATION

Additional supporting information may be found in the online version of this article at http://onlinelibrary.wiley.com/doi/
10.1002/ecy.3226/suppinfo

February 2021 FUNCTIONING OF CORAL REEF COMMUNITIES Article e03226; page 15

http://onlinelibrary.wiley.com/doi/10.1002/ecy.3226/suppinfo

http://onlinelibrary.wiley.com/doi/10.1002/ecy.3226/suppinfo

Journal of Experimental Marine Biology and Ecology 534 (2021) 15147

7

Available online 6 November 2020
0022-0981/© 2020 Elsevier B.V. All rights reserved.

Coral-macroalgal competition under ocean warming and acidification

Lena Rölfer a, b, *, 1, Hauke Reuter a, b, Sebastian C.A. Ferse a, b, Andreas Kubicek c, Sophie Dove c,
Ove Hoegh-Guldberg c, Dorothea Bender-Champ c

a Leibniz Centre for Tropical Marine Research (ZMT), Fahrenheitstraße 6, D-28359 Bremen, Germany
b Faculty of Biology & Chemistry (FB2), University of Bremen, D-28359 Bremen, Germany
c Global Change Institute and School for Biological Sciences, University of Queensland, 4072 Brisbane, Australia

A R T I C L E I N F O

Keywords:
Coral-macroalgal interaction
Ocean acidification
Representative concentration pathways
Porites lobata
Chlorodesmis fastigiata
Climate change

A B S T R A C T

Competition between corals and macroalgae is frequently observed on reefs with the outcome of these in-
teractions affecting the relative abundance of reef organisms and therefore reef health. Anthropogenic activities
have resulted in increased atmospheric CO2 levels and a subsequent rise in ocean temperatures. In addition to
increasing water temperature, elevated CO2 levels are leading to a decrease in oceanic pH (ocean acidification).
These two changes have the potential to alter ecological processes within the oceans, including the outcome of
competitive coral-macroalgal interactions. In our study, we explored the combined effect of temperature increase
and ocean acidification on the competition between the coral Porites lobata and on the Great Barrier Reef
abundant macroalga Chlorodesmis fastigiata. A temperature increase of +1 ◦C above present temperatures and
CO2 increase of +85 ppm were used to simulate a low end emission scenario for the mid- to late 21st century,
according to the Representative Concentration Pathway 2.6 (RCP2.6). Our results revealed that the net photo-
synthesis of P. lobata decreased when it was in contact with C. fastigiata under ambient conditions, and that dark
respiration increased under RCP2.6 conditions. The Photosynthesis to Respiration (P:R) ratios of corals as they
interacted with macroalgal competitors were not significantly different between scenarios. Dark calcification
rates of corals under RCP2.6 conditions, however, were negative and significantly decreased compared to
ambient conditions. Light calcification rates were negatively affected by the interaction of macroalgal contact in
the RCP2.6 scenario, compared to algal mimics and to coral under ambient conditions. Chlorophyll a, and protein
content increased in the RCP2.6 scenario, but were not influenced by contact with the macroalga. We conclude
that the coral host was negatively affected by RCP2.6 conditions, whereas the productivity of its symbionts
(zooxanthellae) was enhanced. While a negative effect of the macroalga (C. fastigiata) on the coral (P. lobata) was
observed for the P:R ratio under control conditions, it was not enhanced under RCP2.6 conditions.

1. Introduction

Macroalgae are important organisms on coral reefs, contributing
significantly to primary production (Gattuso et al., 1998) and nitrogen
fixation (Heil et al., 2004). On a healthy reef, corals generally pre-
dominate the benthic community and are generally competitively su-
perior to macroalgae (Chadwick and Morrow, 2011). However, in recent
years reef ecosystems experienced dramatic declines in coral cover due
to anthropogenic impacts such as global climate change, ocean acidifi-
cation, eutrophication, sedimentation and overfishing as well as disease

outbreaks (Hoegh-Guldberg et al., 2007; Hughes et al., 2010, 2007).
Between 2014 and 2017 a 36 month global heatwave led to multiple
bleaching events on coral reefs and on the Great Barrier Reef to a loss of
shallow water corals of 22–30% with even 50% in the northern parts
(Hughes et al., 2017; Eakin et al., 2018).

The free space on the reef created by high coral mortality can be
taken up by other sessile benthic organisms such as macroalgae, cor-
allimorpharians and sponges (Aronson and Precht, 2001; Norström
et al., 2009). Competition between benthic, sessile organisms is one of
the main factors shaping the community composition on reefs (Dayton,

Abbreviations: RCP2.6, Representative Concentration Pathway 2.6; Pnet, net oxygen production (net photosynthesis); Rdark, dark respiration rate; P:R, Photo-
synthesis per Respiration ratio.

* Corresponding author at: Leibniz Centre for Tropical Marine Research (ZMT), Fahrenheitstraße 6, D-28359 Bremen, Germany.
E-mail address: lena@roelfer.de (L. Rölfer).

1 Present address: Climate Service Center Germany (GERICS), Helmholtz-Zentrum Geesthacht, Fischertwiete 1, D-20095 Hamburg, Germany.

Contents lists available at ScienceDirect

Journal of Experimental Marine Biology and Ecology

journal homepage: www.elsevier.com/locate/jembe

https://doi.org/10.1016/j.jembe.2020.151477
Received 4 June 2020; Received in revised form 2 October 2020; Accepted 29 October 2020

mailto:lena@roelfer.de

www.sciencedirect.com/science/journal/00220981

https://www.elsevier.com/locate/jembe

https://doi.org/10.1016/j.jembe.2020.151477

https://doi.org/10.1016/j.jembe.2020.151477

https://doi.org/10.1016/j.jembe.2020.151477

http://crossmark.crossref.org/dialog/?doi=10.1016/j.jembe.2020.151477&domain=pdf

Journal of Experimental Marine Biology and Ecology 534 (2021) 151477

2

1971). Macroalgae are competing for space in different ways, including
physical abrasion of coral tissue (Coyer et al., 1993), shading (Hughes,
1994) or allelochemicals (i.e. harmful chemicals) to induce bleaching or
death in corals (Longo and Hay, 2017; McCook, 2001; Nugues et al.,
2004). Furthermore, macroalgal exudates can lead to microbe-induced
mortality of adjacent corals (Clements et al., 2020; Smith et al., 2006).
Corals that are weakened by anthropogenic impacts are not able to
invest energy in spatial competition as their energy is needed for various
maintenance functions (Diaz-Pulido et al., 2011; Foster et al., 2008;
Rinkevich and Loya, 1985). Additionally, changes in the reef environ-
ment such as high nutrient availability (eutrophication) and overfishing
of herbivores can result in enhanced growth rates of macroalgae as well
as their release from predation pressure (Hughes et al., 2007, 1987;
Shenkar et al., 2008). As a consequence, macroalgae become the
stronger competitor and proliferate over the reef environment (Done,
1992; McCook, 1999), which may lead to a phase shift from a coral to an
algal dominated state (Anton et al., 2020; Done, 1992; Hughes et al.,
2010; Norström et al., 2009).

Ocean warming and acidification, combined with disturbances such
as overfishing and nutrient enrichment, have a high potential to also
decrease resilience of coral reefs (Anthony et al., 2011; Dove et al.,
2013) and may change the outcome of competition (Chadwick and
Morrow, 2011; Diaz-Pulido et al., 2011; Hoegh-Guldberg et al., 2007).
Thus, on many reefs worldwide macroalgae are the winners in the
competition for space on coral reefs (Gardner, 2003; McCook, 1999;
Mumby et al., 2013; Scheffer et al., 2001).

Corals are particularly vulnerable to ocean acidification (Pörtner
et al., 2019), resulting in a significant reduction in calcification rate
through a decreased aragonite saturation, which controls coral calcifi-
cation (Doney et al., 2009; Kleypas and Langdon, 2006; Langdon, 2002).
With an increase of dissolved CO2, however, productivity may be
enhanced, as CO2 is often a limiting factor in the marine realm.
Enhanced productivity under elevated CO2 levels has been shown for
both, macroalgae (Gao et al., 1993, 1991; Wu et al., 2008) and
zooxanthellae (Al-Moghrabi et al., 1996; Leggat et al., 1999).

Under current conditions, corals are already at their thermal limits,
and with temperatures continuing to rise, corals will be pushed more
frequently beyond their thermal tolerance threshold, as oceans warm
(Hoegh-Guldberg, 1999; Hoegh-Guldberg et al., 2007; Anton et al.,
2020). This was last observed in the longest and most severe coral
bleaching event 2014–2017, when global monthly sea surface temper-
ature maxima increased to 30–31 ◦C (Lough et al., 2018), which led to a
mass mortality of shallow water corals (Heron et al., 2016; Hughes et al.,
2017).

Investigations about coral-algal interactions have strongly increased
during the last decade, monitoring interactions both in situ and in tank
experiments (Bender et al., 2012; Birrell et al., 2008; Brown et al., 2019;
Del Monaco et al., 2017; Diaz-Pulido et al., 2010; Diaz-Pulido and
Barrón, 2020; Jompa and McCook, 2003; McCook, 2001; Rasher et al.,
2011; Rasher and Hay, 2010; Vieira et al., 2016)However, information
on the impacts of global and climate change on ecological interactions
are still underrepresented in the literatura.

In order to explore the interaction of the macroalgae and corals
under climate change, we used the macroalgae Chlorodesmis fastigiata (C.
Agardh) S.C.Ducker and the coral Porites lobata (Dana, 1846) from
Heron Island on the southern Great Barrier Reef. Both species are
abundant and interactions between each are frequently observed
(Jompa and McCook, 2003). C. fastigiata, a siphonous green macroalga,
has been shown to have mostly negative impacts on corals, such as
inducing bleaching (Bonaldo and Hay, 2014), decreased photosynthesis
(Rasher et al., 2011; Rasher and Hay, 2010), polyp retraction by abra-
sion (Jompa and McCook, 2003), reduced tissue recovery (Bender et al.,
2012), and reduced coral settlement (Birrell et al., 2008). P. lobata is a
massive, colonial coral from the order Scleractinia and constitutes one of
the most important and occurring reef-building corals on Pacific coral
reefs (Budd, 1986).

The combined effect of elevated CO2 and temperature on the inter-
action between P. lobata and C. fastigiata was tested against ambient
seawater as control, to provide further insights into coral-algal in-
teractions under changing environmental conditions. Temperature of
the treatment water was increased by 1 ◦C (compared to today; +2 ◦C
compared to preindustrial times) and the CO2-concentration by
+85 ppm according to the RCP2.6 scenario for the mid- to late 21st
century (IPCC, 2013). We hypothesized that the coral would be nega-
tively affected by both, the interaction with macroalgae and RCP2.6
conditions, and that the interaction with live algae would induce
stronger effects than mimics because of biological/chemical in addition
to mechanical effects (e.g. shading or abrasion).

2. Materials and methods

2.1. Study site and collection of corals and macroalgae

This study was done between October and December 2016 (late
spring/Australian summer) at Heron Island Research Station (HIRS),
located in the southern section of the Great Barrier Reef. Organisms
were collected with permission from the Great Barrier Reef Marine Park
Authority (permit number G16/38942.1). Colonies of P. lobata
(approximately 30 cm in diameter) were collected at the reef flat of
Heron Reef (four colonies, 23◦26′S, 151◦54′E) and the adjacent Wistari
Reef (two colonies, 23◦27′S, 151◦52′E). From the colonies, 90 coral
cores (further referred to as corals) of 5 cm diameter were drilled using a
core saw, cut to a height of 1.5 cm and put into tanks with constant flow-
through of seawater for a recovery time of 5 days.

Thirty specimens of C. fastigiata of approximately the same size as
coral cores were collected with a small amount of substrate (Bonaldo
and Hay, 2014) at the reef flat of Heron Reef. Care was taken not to
injure holdfast or other tissue to avoid leaking of chemical compounds.
Macroalgal substrate was cleaned of crabs and other organisms that
lived and fed on the macroalgae were removed. Macroalgae were put
into tanks with corals, with care not to touch the corals (or other mac-
roalgae). After the recovery time of five days, seawater in all tanks was
slowly switched to treatment water over another five days in order to let
corals and algae acclimatize to the new conditions. Macroalgae mimics
were made from 13 cm long pieces of fibre rope (0.7 cm in diameter),
which were bent once in the middle, tied at the base and separated into
its fibres (Edgar and Klumpp, 2003), and pieces of substrate were
attached to the bottom. After the acclimatization time, mimics and
specimens of C. fastigiata were tied to coral cores using rubber bands
(without touching coral tissue) (Bonaldo and Hay, 2014) to avoid
mimics from floating and to assure that macroalgae stayed with the
same coral for the time of the experiment.

2.2. Experimental set-up, maintenance and monitoring

To assess the effect of elevated temperature and ocean acidification
on a coral-algal interaction, organisms were subjected to ambient con-
ditions and the Intergovernmental Panel on Climate Change (IPCC)
scenario RCP2.6 conditions (Dove et al., 2013). Data collected at a
reference site in the Wistari Channel (adjacent to Heron Reef) was used
as a baseline for ambient temperature and CO2. Sea water was pumped
from the reef flat to a holding tower at Heron Island Research Station
and redistributed into two sumps, in which CO2 and temperature
treatment conditions were established (as described by Dove et al.,
2013). From these sumps the experimental tanks were supplied with the
respective treatment water. Temperature and pH feedback sensors in
experimental tanks were connected to a system controller, which then
adjusted the conditions in the sumps at intervals of two hours to guar-
antee exact diurnal and seasonal conditions within the experimental
tanks (Fig. 1). See Dove et al. (2013) for details. Both water treatments
(ambient and RCP2.6) were applied to nine tanks each (18 tanks in
total). Per treatment condition, three tanks each contained five

L. Rölfer et al.

Journal of Experimental Marine Biology and Ecology 534 (2021) 151477

3

replicates of either a) only corals, b) corals with interacting macroalgae,
or c) corals with macroalgal mimics (Fig. 2), summing up to 15 coral
replicates per contact treatment. Organisms were kept under treatment
conditions for a period of 22 days and physiological measurements were
performed subsequently.

Experimental tanks (60*20*38 cm; ~ 35 l) containing corals and
macroalgae were covered with Lee-filters (Old Steel Blue, 725) on the
sides and lids to imitate light intensity at a water depth of 3–6 m on the
reef flat. They were exposed to natural sun irradiance with a constant
flow-through of seawater (0.8–1 l*min− 1). Pumps (Clearpond Infiniti
800) were installed to agitate seawater in order to avoid the forming of a
boundary layer. Tanks were cleaned from fouling organisms every three
days, with macroalgal mimics being washed in freshwater to remove
biofouling, and settled substrate was carefully cleaned off corals with
toothbrushes. At the same time, tanks were rotated to avoid confounding
of light and temperature differences.

To monitor the abiotic conditions within tanks (Table 1), we
installed four light loggers (PAR Sensor, Odyssey, Dataflow Systems, New
Zealand) and five temperature loggers (HOBO Data Loggers, Onset) as
well as four pH-probes (Mettler Toledo, Port Melbourne, Victoria,

Australia, InPro4501VP X connected to an Aquatronica Aquarium
Controller ACQ110). pH was measured on the scale pHsw and the pH
probes were calibrated every second to third day by a two point cali-
bration. The loggers and probes were randomly swapped between tanks
every three days to monitor every tank throughout the experiment.
Temperature and pCO2 contents of sumps were recorded additionally
(Table 1). pCO2 was measured in the sumps (logged continuously every
3 min) and calculated in CO2SYS (developed by E. Lewis and W.R.
Wallace) based on twice daily alkalinity and salinity sampling, and
continuous temperature and pH monitoring at 10 min intervals (see
Dove et al. (2013) for more detail). The total alkalinity of each tank was
sampled once a week at midday and midnight (Table 1) and measured
using a Mettler Toledo titrating system (T50) by Gran titration after
Dickson et al. (2003) using the method with a precision of ±3 μmol Kg− 1
or better as described in Kline et al. (2012). RCP2.6 conditions of ~1 ◦C
increase and a CO2 level of +85 ppm compared to ambient conditions
could be maintained during the experimental period.

Fig. 1. Schematic of sumps and wet table set
up with partial pressure of CO2(pCO2)/
temperature control system. Seawater from
the tower (light blue) is pumped into sumps
(dark blue, red), treatment conditions are
applied and tanks (n = 3 per interaction
treatment) on wet table are connected to
sumps. System controller (grey) with feed-
back loop (dotted line) to adjust conditions
in a two hour interval. AMB: ambient
seawater; RCP2.6: treatment seawater with
increased temperature and pCO2. (For
interpretation of the references to colour in
this figure legend, the reader is referred to
the web version of this article.)

Fig. 2. Pictures of interaction treatments: a) coral P. lobata, b) coral P. lobata with interacting alga C. fastigiata, and c) coral P. lobata with algal mimic (made out of
plastic fibre rope). Ambient and RCP2.6 water treatments were applied to each 3 tanks with 5 replicates each per interaction treatment.

L. Rölfer et al.

Journal of Experimental Marine Biology and Ecology 534 (2021) 151477

4

2.3. Measurement of respiration and calcification rates

In order to investigate treatment effects on the metabolism of corals
and macroalgae, the net oxygen production (Pnet) and dark respiration
rates (Rdark) were measured following the methodology of Crawley et al.
(2010) at the end of the experiment (i.e. after 22 days in treatment
conditions). Organisms (ncoral = 12 per treatment, nmacroalgae = 6 for
RCP2.6 and 8 for ambient) were dark adapted for 45 min prior to dark
respiration measurements. For Pnet measurements the light intensity was
adjusted by modifying the distance of a metal halide lamp (Ocean Light
T5 MH combo 150 W, with 2 × 24 Ocean Blue Actinic, Aqua-Medic of
North America, LLC) from the specimens. The average light intensity
within the experimental tanks matched that measured at midday of the
previous week (694.5 μmol photons m− 2 s− 1 ≈ h = 32.5 cm). Oxygen
flux was recorded using high-precision optical oxygen sensors (optodes)
connected to a logging system (Oxy-10, PreSens, Germany) for 12 and
20 min for macroalgae and corals, respectively, and was normalized to
the surface area of corals (see below) and fresh weight of macroalgae. All
measurements were conducted in the respective treatment water (tam-
bient = 25.9 ◦C, tRCP2.6 = 26.9 ◦C) calculated as a mean for the previous
week. The photosynthesis per respiration (P:R) ratio was calculated by
dividing Pnet by -Rdark.

The light and dark calcification rates were measured in addition to
the photosynthetic/respiratory rates of the corals. To do this, alkalinity
samples (n = 5 per treatment) were taken from respirometry incubations
and the changes in the calcification of corals measured using the alka-
linity anomaly technique using bulk water samples as blanks. Alkalinity
samples were stored in a fridge at 4 ◦C, and within eight weeks after
collection (Huang et al., 2012), were measured using a Mettler Toledo
titrating system (T50) by Gran titration after Dickson et al. (2003) with a
precision of ±3 μmol Kg− 1 or better using the method described in Kline
et al. (2012). The calcification rate was then calculated from the total
alkalinity after Zundelevich et al. (2007), corrected by 10− 3 for unit
conversion and divided by two as molar amount of dissolved CaCO3
equals only half of observed AT increase (Chisholm and Gattuso, 1991).
Calcification rates were normalized to time and surface area and
expressed in μg CaCO3 cm− 2 h− 1.

2.4. Growth measurements

To examine possible differences in growth rates of organisms be-
tween treatments, buoyant weight of corals (Jokiel and Maragos, 1978)
(n = 15 per treatment) and fresh weight of macroalgae (n = 15 per
treatment) were measured before the tank period and after respirometry
measurements were conducted (i.e after 26 days). Fresh weight of
macroalgae was calculated by subtracting the weight of the substrate
(measured by buoyant weight) from the total weight. To minimize in-
accuracy of measurements, macroalgae were blotted with paper tissue
prior to weighing. Weight differences were calculated as percentage
change of buoyant weight and fresh weight for corals and macroalgae,
respectively. After weighing, all samples were frozen at − 20 ◦C pending
further analysis.

2.5. Tissue analysis of corals

Tissue samples were taken from same corals used for respirometry
measurements (n = 12 per treatment), by removal with a seawater jet
(Johannes and Wiebe, 1970). Samples were poured into a Falcon tube
and then vortexed and centrifuged at 4500 rpm for 5 min at 4 ◦C. The
supernatant was poured into a 2 ml Eppendorf tube for protein analysis.
To dilute the pellet left in the tube, 10 ml of filtered seawater was added
and vortexed. From this dilution, 1 ml was pipetted into two tubes for
zooxanthellae count and chlorophyll a analysis.

The population density of zooxanthellae was measured using a
Neubauer hemocytometer (Neubauer-improved, Marienfeld GmbH),
counting cells within three squares of four replicate grids. Chlorophyll a
was measured by adding 2 ml of 100% acetone to the 1 ml pellet dilu-
tion. Samples were sonicated in an ice bath for 10 min to extract the
pigments and then centrifuged at 4500 rpm for 5 min at 4 ◦C to separate
the pellet from the pigment solution. The supernatant was poured into a
new Falcon tube and frozen and the extraction was repeated twice until
supernatant was clear. After the extraction, tubes with supernatants
were centrifuged again to remove debris (Hellebust and Craigie, 1978).
Samples were measured in a spectrophotometer (SpectrostarNano) at
wavelengths of 663 nm and 645 nm and blanks of acetone were
measured after every ten samples and subtracted directly. The amount of
chlorophyll a was then calculated after Arnon (1949) and expressed in g
l− 1. As a quantitative indicator for the thickness of the tissue, the protein
content was measured. Subsamples were measured as triplicates in the
spectrophotometer at wavelengths of 235 nm and 280 nm. Protein
content was calculated after Whitaker and Granum (1980) and
expressed in g l− 1. To normalize tissue properties, respirometry mea-
surements and calcification rates, the corals’ surface area was measured
using the paraffin wax technique (Stimson and Kinzie III, 1991).

2.6. Data analysis

Before statistical analysis of data was performed, all variables were
tested for a possible tank effect by using the lmne package in R Studio,
which compares Gaussian linear and nonlinear mixed-effect models. The
test was performed for models with and without tank as a random factor.
Since the tank factor was non-significant for all variables (p > 0.25)
(results in supplementary table S2), specimens were used as replicates
(Underwood, 1997), hence increasing the power of the analysis.

All response data of corals were tested using a two-factor analysis of
variance (ANOVA) with “scenario” (ambient; RCP2.6) and “contact”
(coral; coral-algal interaction; coral with algal mimic) as fixed factors,
including the interaction term. Response data of macroalgae were tested
using a one-way ANOVA with “scenario” as factor. To account for
multiple comparisons of physiological parameters for P. lobata a Bon-
ferroni correction was applied reducing the α-level of significance to
0.005. When significant effects of factors occurred, ANOVAs were fol-
lowed by a Tukey multiple comparisons test to identify significant
groups. Data were tested for homogeneity of variance (visual inspection

Table 1
Summary of values of water chemistry data for scenario conditions.

Ambient RCP2.6

Temperature [◦C] tanks 25.95 ± 0.65 26.88 ± 0.75
pHSW tanks 8.07 ± 0.03 7.96 ± 0.03
Temperature [◦C] sumps 25.38 ± 0.44 26.29 ± 0.57
pCO2 [ppm] sumps 465.27 ± 53.48 550.58 ± 49.52
Total alkalinity day [μmol Eq. L− 1]

Day 2289.65 ± 12.23 2287.18 ± 11.86
Night 2269.11 ± 26.18 2271.64 ± 25.18

pCO2 [μppm] CO2Sys
Day 268.17 ± 30.32 282.13 ± 25.32
Night 311.26 ± 26.52 331.45 ± 21.68

HCO3
− [μmol Kg SW− 1]

Day 1706.37 ± 43.74 1723.38 ± 36.09
Night 1744.8 ± 35.13 1767.53 ± 31.80

CO3
2− [μmol Kg SW− 1]

Day 234.64 ± 15.85 226.79 ± 11.60
Night 210.61 ± 10.87 202.61 ± 7.98

ΩAragonite
Day 3.66 ± 0.25 3.54 ± 0.18
Night 3.28 ± 0.17 3.16 ± 0.12

Temperature, pH, pCO2 values are means over experimental period, continu-
ously measured over the experimental period (22 days). High standard deviation
is due to daily variability. Total alkalinity, pCO2, HCO3

− , CO3 and ΩAragonite of
treatments are given as means over experimental period, measured once a week
at midday and midnight in tanks and were estimated using CO2SYS software. Kg
SW, kilogram of seawater.

L. Rölfer et al.

Journal of Experimental Marine Biology and Ecology 534 (2021) 151477

5

of residuals vs. fitted values), and normality of residuals was tested using
Shapiro-Wilk normality test. Non-normally distributed data were log or
power transformed to correct for right- or left skew, respectively. Sta-
tistical analysis of data was performed using R Studio Version 1.0.143 (R
Core Team, 2015) and results were expressed as boxplots using the ggplot
package. Some samples of chlorophyll a and protein content of corals
were lost throughout the analysis, reducing the degrees of freedom in
the analysis.

3. Results

3.1. Corals

Coral growth, reported as percentage change in buoyant weight, was
not significantly affected by the RCP2.6 or macroalgae treatment, or a
combination of both. However, slightly negative growth was observed in
the RCP2.6 coral treatment with an average of − 0.34% change in
buoyant weight (see supplementary TableS1).

The technique used for the respirometry measurements allowed the
quantification of calcification rates under treatment conditions in the
presence and absence of light for a particular point in time. The response
of light calcification was dominated by an interaction of “scenario” and
“contact”. In the absence of macroalgae or mimics, corals showed
significantly higher light calcification under ambient conditions with
48.7 ± 10.3 CaCO3 [μg h− 1 cm− 2] as compared to 15.1 ± 5.9 CaCO3
[μg h− 1 cm− 2] in the RCP2.6 scenario (Table 2, Fig. 3a). Corals in the
ambient treatment also performed significantly different compared to
corals in interaction with macroalgae in the RCP2.6 treatment. Dark
calcification was significantly greater in the ambient scenario, with

negative dark calcification in the RCP2.6 scenario for all contact treat-
ments (Table 2, Fig. 3b).

Pnet was significantly influenced by “contact” (Table 2, supplemen-
tary material Fig.S1A) and lower in the coral-algal interaction treatment
compared to only corals, while there was no difference of both treat-
ments to the mimics treatment. Rdark was significantly affected by
“scenario” (Table 2, supplementary material Fig.S1B), with higher
respiration under RCP2.6 compared to ambient conditions. For the P:R
ratio there was a highly significant effect of “contact” (Fig. 3c, Table 2),
with differences among all treatments. P:R ratio was highest for the
mimics treatment, followed by the coral treatment and lowest for the
coral-algal interaction treatment. While the P:R ratio was higher under
ambient conditions for only corals and corals with mimics compared to
RCP2.6 scenario conditions, there was an opposite trend for the coral-
algal interaction. However, the interaction term of “scenario” and
“contact” was not significant (Anova, p = 0.03, Table 2).

Chlorophyll a (Fig. 4d and Protein content (Fig. 4a) were signifi-
cantly different among “scenarios” (Table 2). The mean value of chlo-
rophyll a content in the coral-algal interaction under RCP2.6 conditions
was more than two times higher than for the interaction under ambient
conditions (Fig. 4d). There was a slight trend of a higher zooxanthellae
population density in the RCP2.6 compared to the ambient scenario,
irrespective of the contact treatment (Table 2, Fig. 4b).

3.2. Macroalgae

The macroalga at the centre of this study was very sensitive, and
began to die after eight days of exposure to treatment conditions.
Macroalgae that died were replaced once by ‘back-up’ macroalgae,
which were acclimated and kept in additional tanks with ambient and
RCP2.6 conditions, respectively. However, due to permit limitations no
more macroalgae could be replaced after that. Over the remainder of the
experiment another 14 macroalgae died, some of which were ‘back-up’
macroalgae, summing up to a total of 30 dead macroalgae (n = 13 in
ambient, n = 17 in RCP2.6). Only 16 macroalgae survived throughout
the whole experimental time (n = 9 in ambient, n = 7 in RCP2.6),
reducing the degrees of freedom in the analysis.

Due to differences in size of macroalgae, growth was expressed as
percentage change in fresh weight. Growth of individuals was positive
and negative in both of the scenarios, with negative values resulting
from a loss of filaments, explaining a high standard deviation. However,
total change in fresh weight was positive in both treatments, with no
significant difference between treatments (Table 3). There was also no
significant difference in Rnet between the scenarios (Table 3). Rdark was
slightly higher in the RCP2.6 compared to the ambient scenario,
resulting in a non-significant difference of P:R ratios between scenarios
(Table 3).

4. Discussion

We investigated the important issues as to how ecological competi-
tion may vary under climate change. To do so, we investigated the in-
teractions between the coral P. lobata and the potential competitor, the
fleshy alga C.fastigiate, under low rates of future ocean warming and
acidification. Corals and macroalgae were exposed to a temperature
increase of +1 ◦C and a CO2 increase of +85 ppm above ambient, which
is close to the RCP2.6 scenario of the IPCC at mid- to late 21st century
(IPCC, 2013). This corresponds to CO2 levels expected if action is taken
globally in accordance with the Paris Agreement and refers to the best-
case scenario. We found that both, interaction with macroalgae and
the combined effect of temperature and CO2 affected the coral, whereas
no significant impact of treatment conditions was detected for the
macroalgae.

Table 2
ANOVA output of different variables for P. lobata with bold values indicating
significant effects on the variable.

Variable Source of variation df F p

Bouyant weight Scenario 1 0.279 0.599
Contact 2 1.633 0.202
Scenario x Contact 2 1.587 0.211
Residuals 81

Light calcification Scenario 1 4.354 0.048
Contact 2 2.878 0.076
Scenario x Contact 2 7.480 0.003
Residuals 24

Dark calcification Scenario 1 36.792 <0.001 Contact 2 5.007 0.013 Scenario x Contact 2 2.843 0.078 Residuals 24

Zooxanthellae Scenario 1 7.031 0.010
Contact 2 0.821 0.445
Scenario x Contact 2 0.034 0.967
Residuals 66

Chlorophyll a Scenario 1 13.909 <0.001 Contact 2 0.068 0.934 Scenario x Contact 2 3.434 0.039 Residuals 54

Protein Scenario 1 12.448 <0.001 Contact 2 0.848 0.433 Scenario x Contact 2 2.228 0.117 Residuals 60

Pnet Scenario 1 2.390 0.127
Contact 2 5.893 0.004
Scenario x Contact 2 3.485 0.036
Residuals 66

Rdark Scenario 1 8.749 0.004
Contact 2 3.631 0.032
Scenario x Contact 2 0.156 0.856
Residuals 66

P:R Scenario 1 2.899 0.093
Contact 2 13.466 <0.001 Scenario x Contact 2 3.693 0.030 Residuals 66

df = degrees of freedom; F = F-value; p = p-value (significance <0.005).

L. Rölfer et al.

Journal of Experimental Marine Biology and Ecology 534 (2021) 151477

6

Fig. 3. Calcification rates of P. lobata under a light (n = 5), and b dark conditions (n = 5), measured during respirometry incubations. Letters indicate significant
differences between interactions (a) or among scenarios (b).

Fig. 4. a Protein content (n = 11), b Zooxanthellae density (n = 12), c Photosynthesis/Respiration (P:R) ratio (n = 12), d Chlorophyll a content (n = 10), protein
content (n = 11) for P. lobata. Letters indicate significant differences among scenarios (a,d) or contact (c).

L. Rölfer et al.

Journal of Experimental Marine Biology and Ecology 534 (2021) 151477
7

4.1. Calcification under low rates of warming and acidification

The calcification of massive species such as P. lobata can be slow, as
compared to faster growing corals such as branching species (Lough
et al., 1999). In this study, percentage change in buoyant weight
(~deposited calcium carbonate) was close to zero in all treatments. It is
very likely, however, that the period of our tank experiment (22 days)
was too short, as well as the increase of pCO2 too small, to lead to a
detectable effect in buoyant weight. Anthony et al. (2008) measured
growth of P. lobata over a period of eight weeks and reported slightly
reduced growth rate at 520–700 ppm and a ~ 40% decrease at
1000–1300 ppm. Diaz-Pulido et al. (2011) found that the linear exten-
sion of the fast growing coral Acropora intermedia, measured over eight
weeks, was also strongly negatively affected by CO2 treatments, but
showed no significant difference between treatments with or without
competition with the seaweed Lobophora papenfussii.

While a change in buoyant weight was not detected, measurements
of dark calcification rates were significantly decreased under conditions
similar to RCP2.6 for all interaction treatments, as well as for the coral
treatment under light conditions. This is in agreement with other studies
on coral of the genus Porites, including P. lobata (Anthony et al., 2008);
P. lutea (Ohde and Hossain, 2004); P. compressa (Marubini et al., 2003)
and supports the sensitivity of corals to elevated CO2 (Kleypas and
Langdon, 2006). In the present study, negative calcification (i.e. decal-
cification) was observed in all treatments under RCP2.6 conditions in
the dark. CaCO3 dissolution even exceeded light calcification in the coral
and coral-algae interaction treatment which would lead to negative
growth rates in these treatments if measured over a longer period. While
the amount of energy available from photosynthesis (P:R) was stable
among scenario conditions in the coral-algal interaction, dark calcifi-
cation was reduced under RCP2.6 conditions. This suggests that re-
sources were used for processes other than calcification that demanded
higher energy expenditure under RCP2.6 conditions in the dark while in
contact with the alga.

Calcification under illuminated conditions in the RCP2.6 scenario
was significantly reduced in the interaction with the macroalgae
compared to algal mimics and to coral under ambient conditions. Those
results suggest that macroalgal mimics benefitted the light calcification
of corals through shading by reducing irradiance and therefore light
stress (Anthony et al., 2008), while this positive effect could not be
detected for live algae.

4.2. Photosynthesis and respiration: evidence of a CO2 fertilization effect?

Contrary to the results for the calcification rates, we found that
RCP2.6 conditions significantly increased chlorophyll a and protein
content. The increase in chlorophyll a coincides with an increased net
photosynthesis under RCP2.6 conditions and might be explained by a
‘CO2 fertilization effect’ due to the greater availability of CO2 to
photosynthesize. An increase in chlorophyll an under future conditions
was also found in the branching corals Stylophora pistillata under raised
temperature (Reynaud et al., 2003) and Acropora formosa under elevated
CO2 (Crawley et al., 2010). Crawley et al. (2010) used an increase of

>200 ppm, but interestingly, a CO2 increase of only +85 ppm in this
study was sufficient to lead to an increase in chlorophyll a. This is
comparable to another study, which found an increase in productivity at
an intermediate CO2 scenario (520–705 ppm), while the positive effect
was mitigated at high CO2 (1010–1350 ppm) (Anthony et al., 2008). The
positive effect of CO2 on chlorophyll a found in this study facilitates the
slight increase of zooxanthellae and therefore protein content, which is
an indicator for the nutritional condition of the coral (Ferrier-Pagès
et al., 2003). However, the positive effect of CO2 on the chlorophyll a
content could be mitigated when CO2 concentrations reach a higher
level, as more energy is needed to maintain base functions of the coral
host. This negative effect may be further enhanced by other anthropo-
genic stressors, which weaken the competitive strength of corals over
macroalgae (Diaz-Pulido et al., 2011; Foster et al., 2008).

While a decrease in photosynthesis of corals in contact with various
macroalgae is documented (Rasher et al., 2011; Rasher and Hay, 2010),
the interacting effects of CO2 and temperature on the interaction have
scarcely been considered yet. Our study showed that there was a sig-
nificant decreased P:R ratio of corals in interaction with macroalgae
compared to no contact and the mimics treatment irrespective of the
temperature/CO2 regime the corals were under. RCP2.6 conditions had
a lesser negative effect that was only visible as a trend in the coral only
and mimics treatment. C. fastigiata caused a significant decrease of
photosynthetic efficiency in the corals Montipora digitata, Acropora mil-
lepora and Pocillopora damicornis under ambient conditions, which was
more severe compared to the effect of seven other common macroalgae
(Longo and Hay, 2017; Rasher et al., 2011). In our study, however,
corals were not actually in contact with macroalgae during the physio-
logical measurements, because each species’ metabolic rate was
measured separately. Negative carry-over effects of macroalgae on the
coral photosynthesis and respiration found in our study might be even
more enhanced if measured whilst in contact.

4.3. Physical impacts of competitors

Impacts of macroalgae can harm corals by various mechanisms
including shading and abrasion (McCook et al., 2001) as well as
biochemical reactions, e.g. the induced bleaching due to harmful
chemicals (Longo and Hay, 2017; McCook, 2001; Nugues et al., 2004).
Indeed, C. fastigiata has been shown to produce allelochemicals that can
suppress photosynthesis (Rasher et al., 2011) and cause bleaching
(Rasher and Hay, 2010). A study by Del Monaco et al. (2017) investi-
gated the impact of allelochemical extracts from C. fastigiata on corals
over the same time scale as our project and Diaz-Pulido and Barrón
(2020) tested the release of dissolved organic carbon, which can pro-
mote bacterial metabolism on corals surface and subsequent mortality,
under future CO2 conditions. Both studies conclude that the effects of
C. fastigiata were not more harmful to corals under future climatic
conditions (Del Monaco et al., 2017; Diaz-Pulido and Barrón, 2020).This
is in agreement with our results, which show an effect of C. fastigiata on
the P:R ratio, which however was not enhanced under RCP2.6
conditions.

A recent study by Brown et al. (2019) suggests that coral-algal in-
teractions are temporally variable across seasons. Photosynthetic rates
of the coral A. intermedia in contact with H. heteromorpha were reduced
in winter and increased in summer, while calcification rates in summer
reduced in contact with the algae. Even though photosynthetic activity
was increased in contact with the algae, negative effects of a high-end
ocean acidification and warming scenario (RCP8.5) reduced overall
performance of corals (Brown et al., 2019), which is comparable to the
results of our study.

The impacts of competitors can vary strongly among coral-algal in-
teractions. C. fastigiata is a siphonous macroalga and therefore lacks
discrete cell walls (Rasher et al., 2011). For that reason, handling such as
collection or tank cleaning might have had a significant impact on the
health of the organisms, resulting in a high mortality rate of the

Table 3
ANOVA output of different variables for C. fastigiata.

Variable Source of variation df F p

Fresh weight Scenario 1 1.470 0.244
Residuals 15

Pnet Scenario 1 1.852 0.197
Residuals 13

Rdark Scenario 1 0.336 0.572
Residuals 13

P:R Scenario 1 0.341 0.569
Residuals 13

df = degrees of freedom; F = F-value; p = p-value (significance >0.05).

L. Rölfer et al.

Journal of Experimental Marine Biology and Ecology 534 (2021) 151477

8

macroalgae. Additionally, cycles of periodic loss and reappearance are
known to occur in C. fastigiata, but the timing is unknown (Jompa and
McCook, 2003). As the mortality rate was higher in the RCP2.6 scenario,
temperature and CO2 increase might have affected algal health and led
to a die-off. However, measurements are hardly comparable between
treatments, as time in treatment differed between minimum 13 days and
maximum 22 days, due to the mortality. Furthermore, only 9 corals in
the ambient and 7 corals in the RCP2.6 treatment had an algal partner at
the end of the study. Despite the early die off of macroalgae, we still
measured their effects on corals, which, however, might have been more
visible in the absence of algal mortality, especially in the RCP2.6 sce-
nario, where macroalgae died relatively early in the experiment.

5. Conclusions

P. lobata has a ‘massive ‘coral morphology, and has previously been
shown to be less affected by the interaction with competing macroalgae
compared to other coral species. The negative impact of C. fastigiata was
only visible in the decrease of the P:R ratio, but the study shows no
enhanced impact under RCP2.6 conditions. The energy budget of the
coral in this study, however, was very likely negatively influenced by
RCP2.6 scenario conditions. Calcification, which is directly linked to the
aragonite saturation, was probably negatively affected by the increase in
CO2 as shown in earlier studies. We hypothesize that the productivity of
zooxanthellae might be enhanced under the RCP2.6 scenario due to
elevated CO2 availability (CO2 fertilization effect), leading to an in-
crease of chlorophyll a. The coral host, however, was rather stressed,
resulting in higher respiration and decreased calcification.

C. fastigiata is known for its strong allelopathy, but also very sensitive
under experimental conditions. While the impacts of the algae on the
coral were small, a temperature and CO2 increase of more than 1 ◦C and
85 ppm respectively over longer periods, whether due to global warming
or warm water periods (e.g. El Niño), might have significant impacts on
coral-macroalgal interactions. Hence, further studies with less sensitive
macroalgae are needed to investigate the likelihood of interaction shifts
for P. lobata under future climatic regimes.

Declaration of Competing Interest

The authors declare that they have no known competing financial
interests or personal relationships that could have appeared to influence
the work reported in this paper.

Acknowledgements

We would like to thank the University of Queensland and the Leibniz
Centre for Tropical Marine Research (ZMT) for funding. LR was sup-
ported by the ZMT ‘Student Research Grant’. DB-C and OHG were
supported by the Australian Research Council Laureate Fellowship
program. Thanks to the research assistants at Heron Island Research
Station for technical support.

Appendix A. Supplementary data

Supplementary data to this article can be found online at https://doi.
org/10.1016/j.jembe.2020.151477.

References

Al-Moghrabi, S., Goiran, C., Allemand, D., Speziale, N., Jaubert, J., 1996. Inorganic
carbon uptake for photosynthesis by the symbiotic coral-dinoflagellate association II.
Mechanisms for bicarbonate uptake. J. Exp. Mar. Biol. Ecol. 199, 227–248. https://
doi.org/10.1016/0022-0981(95)00202-2.

Anthony, Kline, D.I., Diaz-Pulido, G., Dove, S., Hoegh-Guldberg, O., 2008. Ocean
acidification causes bleaching and productivity loss in coral reef builders. Proc. Natl.
Acad. Sci. 105, 17442–17446. https://doi.org/10.1073/pnas.0804478105.

Anthony, K.R., Maynard, J.A., Diaz-Pulido, G., Mumby, P.J., Marshall, P.A., Cao, L.,
Hoegh-Guldberg, O., 2011. Ocean acidification and warming will lower coral reef
resilience. Glob. Chang. Biol. 17, 1798–1808.

Anton, A., Randle, J.L., Garcia, F.C., Rossbach, S., Ellis, J.I., Weinzierl, M., Duarte, C.M.,
2020. Differential thermal tolerance between algae and corals may trigger the
proliferation of algae in coral reefs. Glob. Chang. Biol. 26, 4316–4327. https://doi.
org/10.1111/gcb.15141.

Arnon, D.I., 1949. Copper enzymes in isolated chloroplasts. Polyphenoloxidase in Beta
vulgaris. Plant Physiol. 24 (1).

Aronson, R.B., Precht, W.F., 2001. White-band disease and the changing face of
Caribbean coral reefs. In: Porter, J.W. (Ed.), The Ecology and Etiology of Newly
Emerging Marine Diseases. Springer Netherlands, Dordrecht, pp. 25–38. https://doi.
org/10.1007/978-94-017-3284-0_2.

Bender, D., Diaz-Pulido, G., Dove, S., 2012. Effects of macroalgae on corals recovering
from disturbance. J. Exp. Mar. Biol. Ecol. 429, 15–19. https://doi.org/10.1016/j.
jembe.2012.06.014.

Birrell, C.L., McCook, L.J., Willis, B.L., Diaz-Pulido, G.A., 2008. Effects of benthic algae
on the replenishment of corals and the implications for the resilience of coral reefs.
Ocean. Mar. Biol. Annu. Rev. 46, 25–63.

Bonaldo, R.M., Hay, M.E., 2014. Seaweed-coral interactions: variance in seaweed
allelopathy, coral susceptibility, and potential effects on coral resilience. PLoS One 9,
e85786.

Budd, A.F., 1986. Neogene Paleontology in the Northern Dominican Republic: 3. The
Family Poritidae (Anthozoa: Scleractinia).

Brown, K.T., Bender-Champ, D., Kenyon, T.M., Rémond, C., Hoegh-Guldberg, O.,
Dove, S., 2019. Temporal effects of ocean warming and acidification on coral–algal
competition. Coral Reefs 38, 297–309. https://doi.org/10.1007/s00338-019-01775-
y.

Chadwick, N.E., Morrow, K.M., 2011. Competition among sessile organisms on coral
reefs. In: Dubinsky, Z., Stambler, N. (Eds.), Coral Reefs: An Ecosystem in Transition.
Springer Netherlands, Dordrecht, pp. 347–371. https://doi.org/10.1007/978-94-
007-0114-4_20.

Chisholm, J.R., Gattuso, J.-P., 1991. Validation of the alkalinity anomaly technique for
investigating calcification of photosynthesis in coral reef communities. Limnol.
Oceanogr. 36, 1232–1239.

Clements, C.S., Burns, A.S., Stewart, F.J., Hay, M.E., 2020. Seaweed-coral competition in
the field: effects on coral growth, photosynthesis and microbiomes require direct
contact. Proc. Biol. Sci. 287, 20200366 https://doi.org/10.1098/rspb.2020.0366.

Coyer, J.A., Ambrose, R.F., Engle, J.M., Carroll, J.C., 1993. Interactions between corals
and algae on a temperate zone rocky reef: mediation by sea urchins. J. Exp. Mar.
Biol. Ecol. 167, 21–37.

Crawley, A., Kline, D.I., Dunn, S., Anthony, K., Dove, S., 2010. The effect of ocean
acidification on symbiont photorespiration and productivity in Acropora formosa.
Glob. Chang. Biol. 16, 851–863. https://doi.org/10.1111/j.1365-2486.2009.01943.
x.

Dayton, P.K., 1971. Competition, disturbance, and community organization: the
provision and subsequent utilization of space in a rocky intertidal community. Ecol.
Monogr. 41, 351–389. https://doi.org/10.2307/1948498.

Del Monaco, C., Hay, M.E., Gartrell, P., Mumby, P.J., Diaz-Pulido, G., 2017. Effects of
ocean acidification on the potency of macroalgal allelopathy to a common coral. Sci.
Rep. 7, 41053 https://doi.org/10.1038/srep41053.

Diaz-Pulido, G., Barrón, C., 2020. CO2 enrichment stimulates dissolved organic carbon
release in coral reef macroalgae. J. Phycol. 56, 1039–1052. https://doi.org/
10.1111/jpy.13002.

Diaz-Pulido, G., Harii, S., McCook, L.J., Hoegh-Guldberg, O., 2010. The impact of benthic
algae on the settlement of a reef-building coral. Coral Reefs 29, 203–208. https://
doi.org/10.1007/s00338-009-0573-x.

Diaz-Pulido, G., Gouezo, M., Tilbrook, B., Dove, S., Anthony, K.R.N., 2011. High CO2
enhances the competitive strength of seaweeds over corals: coral-algal competition
and high CO2. Ecol. Lett. 14, 156–162. https://doi.org/10.1111/j.1461-
0248.2010.01565.x.

Dickson, A.G., Afghan, J.D., Anderson, G.C., 2003. Reference materials for oceanic CO2
analysis: a method for the certification of total alkalinity. Mar. Chem. 80, 185–197.

Done, T.J., 1992. Phase shifts in coral reef communities and their ecological significance.
Hydrobiologia 247, 121–132. https://doi.org/10.1007/BF00008211.

Doney, S.C., Fabry, V.J., Feely, R.A., Kleypas, J.A., 2009. Ocean acidification: the other
CO2 problem. Annu. Rev. Mar. Sci. 1, 169–192. https://doi.org/10.1146/annurev.
marine.010908.163834.

Dove, S.G., Kline, D.I., Pantos, O., Angly, F.E., Tyson, G.W., Hoegh-Guldberg, O., 2013.
Future reef decalcification under a business-as-usual CO2 emission scenario. Proc.
Natl. Acad. Sci. 110, 15342–15347. https://doi.org/10.1073/pnas.1302701110.

Eakin, C.M., Liu, G., Gomez, A.M., De la Couri, J.L., Heron, S.F., Skirving, W.J., Geiger, E.
F., Marsh, B.L., Tirak, K.V., Strong, A.E., 2018. Unprecedented three years of global
coral bleaching 2014–17. Sidebar 3.1. [in State of the Climate in 2017]. Bull. Am.
Meteorol. Soc. 99 (8), S74–S75.

Edgar, G.J., Klumpp, D.W., 2003. Consistencies over regional scales in assemblages of
mobile epifauna associated with natural and artificial plants of different shape.
Aquat. Bot. 75, 275–291.

Ferrier-Pagès, C., Witting, J., Tambutté, E., Sebens, K.P., 2003. Effect of natural
zooplankton feeding on the tissue and skeletal growth of the scleractinian coral
Stylophora pistillata. Coral Reefs 22, 229–240. https://doi.org/10.1007/s00338-
003-0312-7.

Foster, N., Box, S., Mumby, P., 2008. Competitive effects of macroalgae on the fecundity
of the reef-building coral Montastraea annularis. Mar. Ecol. Prog. Ser. 367, 143–152.
https://doi.org/10.3354/meps07594.

L. Rölfer et al.

https://doi.org/10.1016/j.jembe.2020.151477

https://doi.org/10.1016/j.jembe.2020.151477

https://doi.org/10.1016/0022-0981(95)00202-2

https://doi.org/10.1016/0022-0981(95)00202-2

https://doi.org/10.1073/pnas.0804478105

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0015

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0015

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0015

https://doi.org/10.1111/gcb.15141

https://doi.org/10.1111/gcb.15141

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0025

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0025

https://doi.org/10.1007/978-94-017-3284-0_2

https://doi.org/10.1007/978-94-017-3284-0_2

https://doi.org/10.1016/j.jembe.2012.06.014

https://doi.org/10.1016/j.jembe.2012.06.014

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0040

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0040

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0040

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0045

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0045

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0045

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0050

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0050

https://doi.org/10.1007/s00338-019-01775-y

https://doi.org/10.1007/s00338-019-01775-y

https://doi.org/10.1007/978-94-007-0114-4_20

https://doi.org/10.1007/978-94-007-0114-4_20

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0060

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0060

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0060

https://doi.org/10.1098/rspb.2020.0366

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0070

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0070

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0070

https://doi.org/10.1111/j.1365-2486.2009.01943.x

https://doi.org/10.1111/j.1365-2486.2009.01943.x

https://doi.org/10.2307/1948498

https://doi.org/10.1038/srep41053

https://doi.org/10.1111/jpy.13002

https://doi.org/10.1111/jpy.13002

https://doi.org/10.1007/s00338-009-0573-x

https://doi.org/10.1007/s00338-009-0573-x

https://doi.org/10.1111/j.1461-0248.2010.01565.x

https://doi.org/10.1111/j.1461-0248.2010.01565.x

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0110

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0110

https://doi.org/10.1007/BF00008211

https://doi.org/10.1146/annurev.marine.010908.163834

https://doi.org/10.1146/annurev.marine.010908.163834

https://doi.org/10.1073/pnas.1302701110

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0130

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0130

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0130

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0130

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0135

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0135

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0135

https://doi.org/10.1007/s00338-003-0312-7

https://doi.org/10.1007/s00338-003-0312-7

https://doi.org/10.3354/meps07594

Journal of Experimental Marine Biology and Ecology 534 (2021) 151477

9

Gao, K., Aruga, Y., Asada, K., Ishihara, T., Akano, T., Kiyohara, M., 1991. Enhanced
growth of the red algaPorphyra yezoensis Ueda in high CO2 concentrations. J. Appl.
Phycol. 3, 355–362.

Gao, K., Aruga, Y., Asada, K., Kiyohara, M., 1993. Influence of enhanced CO2 on growth
and photosynthesis of the red algae Gracilaria sp. and G. chilensis. J. Appl. Phycol. 5,
563–571. https://doi.org/10.1007/BF02184635.

Gardner, T.A., 2003. Long-term region-wide declines in Caribbean corals. Science 301,
958–960. https://doi.org/10.1126/science.1086050.

Gattuso, J.-P., Frankignoulle, M., Wollast, R., 1998. Carbon and carbonate metabolism in
coastal aquatic ecosystems. Annu. Rev. Ecol. Syst. 29, 405–434.

Heil, C.A., Chaston, K., Jones, A., Bird, P., Longstaff, B., Costanzo, S., Dennison, W.C.,
2004. Benthic microalgae in coral reef sediments of the southern great barrier reef,
Australia. Coral Reefs 23, 336–343. https://doi.org/10.1007/s00338-004-0390-1.

Hellebust, J.A., Craigie, J.S., 1978. Handbook of Phycological Methods Physiological and
Biological Methods. Cambridge University Press Cambridge, London.

Heron, S.F., Maynard, J.A., van Hooidonk, R., Eakin, C.M., 2016. Warming trends and
bleaching stress of the World’s coral reefs 1985–2012. Sci. Rep. 6 https://doi.org/
10.1038/srep38402.

Hoegh-Guldberg, O., 1999. Climate change, coral bleaching and the future of the World’s
coral reefs. Mar. Freshw. Res. 50, 839. https://doi.org/10.1071/MF99078.

Hoegh-Guldberg, O., Mumby, P.J., Hooten, A.J., Steneck, R.S., Greenfield, P., Gomez, E.,
Harvell, C.D., Sale, P.F., Edwards, A.J., Caldeira, K., et al., 2007. Coral reefs under
rapid climate change and ocean acidification. Science 318, 1737–1742.

Huang, W.-J., Wang, Y., Cai, W., 2012. Assessment of sample storage techniques for total
alkalinity and dissolved inorganic carbon in seawater. Limnol. Oceanogr. Methods
10. https://doi.org/10.4319/lom.2012.10.711.

Hughes, T.P., 1994. Catastrophes, phase shifts, and large-scale degradation of a
Caribbean coral reef. Science 265, 1547–1551. https://doi.org/10.1126/
science.265.5178.1547.

Hughes, T.P., Reed, D.C., Boyle, M.-J., 1987. Herbivory on coral reefs: community
structure following mass mortalities of sea urchins. J. Exp. Mar. Biol. Ecol. 113,
39–59. https://doi.org/10.1016/0022-0981(87)90081-5.

Hughes, T.P., Rodrigues, M.J., Bellwood, D.R., Ceccarelli, D., Hoegh-Guldberg, O.,
McCook, L., Moltschaniwskyj, N., Pratchett, M.S., Steneck, R.S., Willis, B., 2007.
Phase shifts, herbivory, and the resilience of coral reefs to climate change. Curr. Biol.
17, 360–365. https://doi.org/10.1016/j.cub.2006.12.049.

Hughes, T.P., Graham, N.A.J., Jackson, J.B.C., Mumby, P.J., Steneck, R.S., 2010. Rising
to the challenge of sustaining coral reef resilience. Trends Ecol. Evol. 25, 633–642.
https://doi.org/10.1016/j.tree.2010.07.011.

Hughes, T.P., Kerry, J.T., Álvarez-Noriega, M., Álvarez-Romero, J.G., Anderson, K.D.,
Baird, A.H., Babcock, R.C., Beger, M., Bellwood, D.R., Berkelmans, R., Bridge, T.C.,
Butler, I.R., Byrne, M., Cantin, N.E., Comeau, S., Connolly, S.R., Cumming, G.S.,
Dalton, S.J., Diaz-Pulido, G., Eakin, C.M., Figueira, W.F., Gilmour, J.P., Harrison, H.
B., Heron, S.F., Hoey, A.S., Hobbs, J.-P.A., Hoogenboom, M.O., Kennedy, E.V.,
Kuo, C., Lough, J.M., Lowe, R.J., Liu, G., McCulloch, M.T., Malcolm, H.A.,
McWilliam, M.J., Pandolfi, J.M., Pears, R.J., Pratchett, M.S., Schoepf, V.,
Simpson, T., Skirving, W.J., Sommer, B., Torda, G., Wachenfeld, D.R., Willis, B.L.,
Wilson, S.K., 2017. Global warming and recurrent mass bleaching of corals. Nature
543, 373–377. https://doi.org/10.1038/nature21707.

IPCC, 2013. Climate Change 2013 – The Physical Science Basis, Climate Change 2013:
The Physical Science Basis. Contribution of Working Group I to the Fifth Assessment
Report of the Intergovernmental Panel on Climate Change. Cambridge University
Press, Cambridge, United Kingdom and New York, NY, USA. https://doi.org/
10.1017/cbo9781107415324.

Johannes, R.E., Wiebe, W.J., 1970. Method for determination of coral tissue biomass and
composition. Limnol. Oceanogr. 15, 822–824.

Jokiel, P.L., Maragos, J.E., 1978. Coral growth: Buoyant weight technique. In:
Stoddart, D.R., Johannes, R.E. (Eds.), Coral Reefs: Research Methods, Monographs
on Oceanographic Methodology. UNESCO, Paris, pp. 529–542.

Jompa, J., McCook, L.J., 2003. Contrasting effects of turf algae on corals: massive Porites
spp. are unaffected by mixed-species turfs, but killed by the red alga Anotrichium
tenue. Mar. Ecol. Prog. Ser. 258, 79–86.

Kleypas, J.A., Langdon, C., 2006. Coral reefs and changing seawater carbonate
chemistry. In: Phinney, J.T., Hoegh-Guldberg, O., Kleypas, J., Skirving, W.,
Strong, A. (Eds.), Coastal and Estuarine Studies. American Geophysical Union,
Washington, D. C, pp. 73–110. https://doi.org/10.1029/61CE06.

Kline, D.I., Teneva, L., Schneider, K., Miard, T., Chai, A., Marker, M., Headley, K.,
Opdyke, B., Nash, M., Valetich, M., Caves, J.K., Russell, B.D., Connell, S.D.,
Kirkwood, B.J., Brewer, P., Peltzer, E., Silverman, J., Caldeira, K., Dunbar, R.B.,
Koseff, J.R., Monismith, S.G., Mitchell, B.G., Dove, S., Hoegh-Guldberg, O., 2012.
A short-term in situ CO2 enrichment experiment on Heron Island (GBR). Sci. Rep. 2
https://doi.org/10.1038/srep00413.

Langdon, C., 2002. Review of experimental evidence for effects of CO2 on calcification of
reef builders. In: Proc. 9th Int. Coral Reef Sym, pp. 1091–1098.

Leggat, W., Badger, M.R., Yellowlees, D., 1999. Evidence for an inorganic carbon-
concentrating mechanism in the symbiotic Dinoflagellate Symbiodinium sp. Plant
Physiol. 121, 1247–1255. https://doi.org/10.1104/pp.121.4.1247.

Longo, G.O., Hay, M.E., 2017. Seaweed allelopathy to corals: are active compounds on,
or in, seaweeds? Coral Reefs 36, 247–253. https://doi.org/10.1007/s00338-016-
1526-9.

Lough, J.M., Barnes, D.J., Devereux, M.J., Tobin, B.J., Tobin, S., 1999. Variability in
growth characteristics of massive Porites on the great barrier reef. CRC Reef Res.
Cent. Tech. Rep. 95.

Lough, J.M., Anderson, K.D., Hughes, T.P., 2018. Increasing thermal stress for tropical
coral reefs: 1871–2017. Sci. Rep. 8, 6079. https://doi.org/10.1038/s41598-018-
24530-9.

Marubini, F., Ferrier-Pages, C., Cuif, J.-P., 2003. Suppression of skeletal growth in
scleractinian corals by decreasing ambient carbonate-ion concentration: a cross-
family comparison. Proc. R. Soc. Lond. B Biol. Sci. 270, 179–184.

McCook, L.J., 1999. Macroalgae, nutrients and phase shifts on coral reefs: scientific
issues and management consequences for the great barrier reef. Coral Reefs 18,
357–367.

McCook, L., 2001. Competition between corals and algal turfs along a gradient of
terrestrial influence in the nearshore central great barrier reef. Coral Reefs 19,
419–425. https://doi.org/10.1007/s003380000119.

McCook, L., Jompa, J., Diaz-Pulido, G., 2001. Competition between corals and algae on
coral reefs: a review of evidence and mechanisms. Coral Reefs 19, 400–417. https://
doi.org/10.1007/s003380000129.

Mumby, P.J., Steneck, R.S., Hastings, A., 2013. Evidence for and against the existence of
alternate attractors on coral reefs. Oikos 122, 481–491. https://doi.org/10.1111/
j.1600-0706.2012.00262.x.

Norström, A., Nyström, M., Lokrantz, J., Folke, C., 2009. Alternative states on coral reefs:
beyond coral–macroalgal phase shifts. Mar. Ecol. Prog. Ser. 376, 295–306. https://
doi.org/10.3354/meps07815.

Nugues, M.M., Smith, G.W., Hooidonk, R.J., Seabra, M.I., Bak, R.P.M., 2004. Algal
contact as a trigger for coral disease: algal contact triggers coral disease. Ecol. Lett. 7,
919–923. https://doi.org/10.1111/j.1461-0248.2004.00651.x.

Ohde, S., Hossain, M.M.M., 2004. Effect of CaCO3 (aragonite) saturation state of
seawater on calcification of Porites coral. Geochem. J. 38, 613–621.

Pörtner, H.O., Roberts, D.C., Masson-Delmotte, V., Zhai, P., Tignor, M., Poloczanska, E.,
Mintebeck, K., Nicolai, M., Okem, A., Petzold, J., Rama, B., Weyer, N., 2019. IPCC
The Ocean and Cryosphere in a Changing Climate Summary for Policmakers. IPCC
Spec. Rep. Ocean Cryosph. a Chang. Clim. SPM-1-SPM-42.

R Core Team, 2015. R: A Language and Environment for Statistical Computing. R
Foundation for Statistical Computing, Vienna, Austria.

Rasher, D.B., Hay, M.E., 2010. Chemically rich seaweeds poison corals when not
controlled by herbivores. Proc. Natl. Acad. Sci. 107, 9683–9688. https://doi.org/
10.1073/pnas.0912095107.

Rasher, D.B., Stout, E.P., Engel, S., Kubanek, J., Hay, M.E., 2011. Macroalgal terpenes
function as allelopathic agents against reef corals. Proc. Natl. Acad. Sci. 108,
17726–17731. https://doi.org/10.1073/pnas.1108628108.

Reynaud, S., Leclercq, N., Romaine-Lioud, S., Ferrier-Pages, C., Jaubert, J., Gattuso, J.-P.,
2003. Interacting effects of CO2 partial pressure and temperature on photosynthesis
and calcification in a scleractinian coral. Glob. Chang. Biol. 9, 1660–1668. https://
doi.org/10.1046/j.1365-2486.2003.00678.x.

Rinkevich, B., Loya, Y., 1985. Intraspecific competition in a reef coral: effects on growth
and reproduction. Oecologia 66, 100–105. https://doi.org/10.1007/BF00378559.

Scheffer, M., Carpenter, S., Foley, J.A., Folke, C., Walker, B., 2001. Catastrophic shifts in
ecosystems. Nature 413, 591–596. https://doi.org/10.1038/35098000.

Shenkar, N., Bronstein, O., Loya, Y., 2008. Population dynamics of a coral reef ascidian
in a deteriorating environment. Mar. Ecol. Prog. Ser. 367, 163–171. https://doi.org/
10.3354/meps07579.

Smith, J.E., Shaw, M., Edwards, R.A., Obura, D., Pantos, O., Sala, E., Sandin, S.A.,
Smriga, S., Hatay, M., Rohwer, F.L., 2006. Indirect effects of algae on coral: Algae-
mediated, microbe-induced coral mortality. Ecol. Lett. 9, 835–845. https://doi.org/
10.1111/j.1461-0248.2006.00937.x.

Stimson, J., Kinzie III, R.A., 1991. The temporal pattern and rate of release of
zooxanthellae from the reef coral Pocillopora damicornis (Linnaeus) under nitrogen-
enrichment and control conditions. J. Exp. Mar. Biol. Ecol. 153, 63–74.

Underwood, A.J., 1997. Experiments in Ecology: Their Logical Design and Interpretation
Using Analysis of Variance. Cambridge University Press.

Vieira, C., Thomas, O.P., Culioli, G., Genta-Jouve, G., Houlbreque, F., Gaubert, J., De
Clerck, O., Payri, C.E., 2016. Allelopathic interactions between the brown algal
genus Lobophora (Dictyotales, Phaeophyceae) and scleractinian corals. Sci. Rep. 6,
18637.

Whitaker, J.R., Granum, P.E., 1980. An absolute method for protein determination based
on difference in absorbance at 235 and 280 nm. Anal. Biochem. 109, 156–159.

Wu, H., Zou, D., Gao, K., 2008. Impacts of increased atmospheric CO2 concentration on
photosynthesis and growth of micro- and macro-algae. Sci China C Life Sci 51,
1144–1150. https://doi.org/10.1007/s11427-008-0142-5.

Zundelevich, A., Lazar, B., Ilan, M., 2007. Chemical versus mechanical bioerosion of
coral reefs by boring sponges-lessons from Pione cf. vastifica. J. Exp. Biol. 210,
91–96.

L. Rölfer et al.

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0150

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0150

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0150

https://doi.org/10.1007/BF02184635

https://doi.org/10.1126/science.1086050

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0165

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0165

https://doi.org/10.1007/s00338-004-0390-1

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0175

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0175

https://doi.org/10.1038/srep38402

https://doi.org/10.1038/srep38402

https://doi.org/10.1071/MF99078

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0190

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0190

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0190

https://doi.org/10.4319/lom.2012.10.711

https://doi.org/10.1126/science.265.5178.1547

https://doi.org/10.1126/science.265.5178.1547

https://doi.org/10.1016/0022-0981(87)90081-5

https://doi.org/10.1016/j.cub.2006.12.049

https://doi.org/10.1016/j.tree.2010.07.011

https://doi.org/10.1038/nature21707

https://doi.org/10.1017/cbo9781107415324

https://doi.org/10.1017/cbo9781107415324

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0225

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0225

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0230

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0230

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0230

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0235

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0235

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0235

https://doi.org/10.1029/61CE06

https://doi.org/10.1038/srep00413

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0250

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0250

https://doi.org/10.1104/pp.121.4.1247

https://doi.org/10.1007/s00338-016-1526-9

https://doi.org/10.1007/s00338-016-1526-9

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0265

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0265

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0265

https://doi.org/10.1038/s41598-018-24530-9

https://doi.org/10.1038/s41598-018-24530-9

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0275

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0275

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0275

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0280

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0280

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0280

https://doi.org/10.1007/s003380000119

https://doi.org/10.1007/s003380000129

https://doi.org/10.1007/s003380000129

https://doi.org/10.1111/j.1600-0706.2012.00262.x

https://doi.org/10.1111/j.1600-0706.2012.00262.x

https://doi.org/10.3354/meps07815

https://doi.org/10.3354/meps07815

https://doi.org/10.1111/j.1461-0248.2004.00651.x

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0310

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0310

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0315

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0315

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0315

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0315

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0320

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0320

https://doi.org/10.1073/pnas.0912095107

https://doi.org/10.1073/pnas.0912095107

https://doi.org/10.1073/pnas.1108628108

https://doi.org/10.1046/j.1365-2486.2003.00678.x

https://doi.org/10.1046/j.1365-2486.2003.00678.x

https://doi.org/10.1007/BF00378559

https://doi.org/10.1038/35098000

https://doi.org/10.3354/meps07579

https://doi.org/10.3354/meps07579

https://doi.org/10.1111/j.1461-0248.2006.00937.x

https://doi.org/10.1111/j.1461-0248.2006.00937.x

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0355

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0355

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0355

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0360

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0360

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0365

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0365

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0365

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0365

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0370

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0370

https://doi.org/10.1007/s11427-008-0142-5

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0380

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0380

http://refhub.elsevier.com/S0022-0981(20)30246-X/rf0380

Coral-macroalgal competition under ocean warming and acidification
1 Introduction
2 Materials and methods
2.1 Study site and collection of corals and macroalgae
2.2 Experimental set-up, maintenance and monitoring
2.3 Measurement of respiration and calcification rates
2.4 Growth measurements
2.5 Tissue analysis of corals
2.6 Data analysis
3 Results
3.1 Corals
3.2 Macroalgae
4 Discussion
4.1 Calcification under low rates of warming and acidification
4.2 Photosynthesis and respiration: evidence of a CO2 fertilization effect?
4.3 Physical impacts of competitors
5 Conclusions
Declaration of Competing Interest
Acknowledgements
Appendix A Supplementary data
References

Journal of Experimental Marine Biology and Ecology 535 (2021) 15148

9

Available online 13 November 2020
0022-0981/© 2020 Elsevier B.V. All rights reserved.

Irradiance, photosynthesis and elevated pCO2 effects on net calcification in
tropical reef macroalgae

C. McNicholl , M.S. Koch *

Florida Atlantic University, Boca Raton, FL 33431, USA

A R T I C L E I N F O

Keywords:
Coral reef
Dissolution
pH
Climate change
Ocean acidification

A B S T R A C T

Calcifying tropical macroalgae produce sediment, build three-dimensional habitats, and provide substrate for
invertebrate larvae on reefs. Thus, lower calcification rates under declining pH and increasing ocean pCO2, or
ocean acidification, is a concern. In the present study, calcification rates were examined experimentally under
predicted end-of-the-century seawater pCO2 (1116 μatm) and pH (7.67) compared to ambient controls (pCO2
409 μatm; pH 8.04). Nine reef macroalgae with diverse calcification locations, calcium carbonate structure,
photophysiology, and site-specific irradiance were examined under light and dark conditions. Species included
five from a high light patch reef on the Florida Keys Reef Tract (FKRT) and four species from low light reef walls
on Little Cayman Island (LCI). Experiments on FKRT and LCI species were conducted at 500 and 50 μmol photons
m− 2 s− 1 in situ irradiance, respectively. Calcification rates independent of photosystem-II (PSII) were also
investigated for FKRT species. The most consistent negative effect of elevated pCO2 on calcification rates in the
tropical macroalgae examined occurred in the dark. Most species (89%) had net calcification rates of zero or net
dissolution in the dark at low pH. Species from the FKRT that sustained positive net calcification rates in the light
at low pH also maintained ~30% of their net calcification rates without PSII at ambient pH. However, calcifi-
cation rates in the light independent of PSII were not sustained at low pH. Regardless of these low pH effects,
most FKRT species daily net calcification rates, integrating light/dark rates over a 24h period, were not signif-
icantly different between low and ambient pH. This was due to a 10-fold lower dark, compared to light, calci-
fication rate, and a strong correspondence between calcification and photosynthetic rates. Interestingly, low-light
species sustained calcification rates on par with high-light species without high rates of photosynthesis. Low-light
species’ morphology and physiology that promote high calcification rates at ambient pH, may increase their
vulnerability to low pH. Our data indicate that the negative effect of elevated pCO2 and low pH on tropical
macroalgae at the organismal level is their impact on dark net calcification, probably enhanced dissolution.
However, elevated pCO2 and low pH effects on macroalgae daily calcification rates are greatest in species with
lower net calcification rates in the light. Thus, macroalgae able to maintain high calcification rates in the light
(high and low irradiance) at low pH, and/or sustain strong biotic control with high [H+] in the bulk seawater, are
expected to dominate under global change.

1. Introduction

Marine calcifier persistence and sustained calcification rates remain
uncertain under future predictions of ocean acidification. Since the in-
dustrial revolution, global ocean pH has decreased 0.1 pH units, and a
further 0.3–0.4 reduction is predicted to occur by the year 2100 due to
anthropogenic CO2 emissions (Gehlen et al. 2014; Hartin et al. 2016). A
global decrease in ocean pH affects calcification rates in marine organ-
isms, such as coral, shellfish, phytoplankton, and ecologically important

macroalgae (Andersson et al. 2009; Fabry et al. 2008; Hoegh-Guldberg
et al. 2007; Koch et al. 2013; Orr et al. 2005; Ries et al. 2009). Dimin-
ished calcification rates of tropical reef macroalgae is a concern because
of their ecological role in carbonate sediment production, building of 3-
dimensional reef habitat structure, and providing substrate for inverte-
brate larval settlement (Adey 1998; Nelson 2009). Many studies have
examined the effects of elevated partial pressure of CO2 (pCO2) that
lowers seawater pH on macroalgal calcification, but often with con-
flicting results (Hofmann et al. 2014; Koch et al. 2013; Nelson 2009;

* Corresponding author.
E-mail addresses: cmcnicholl2015@fau.edu (C. McNicholl), mkoch@fau.edu (M.S. Koch).

Contents lists available at ScienceDirect

Journal of Experimental Marine Biology and Ecology

journal homepage: www.elsevier.com/locate/jembe

https://doi.org/10.1016/j.jembe.2020.151489
Received 19 November 2019; Received in revised form 22 October 2020; Accepted 2 November 2020

mailto:cmcnicholl2015@fau.edu

mailto:mkoch@fau.edu

www.sciencedirect.com/science/journal/00220981

https://www.elsevier.com/locate/jembe

https://doi.org/10.1016/j.jembe.2020.151489

https://doi.org/10.1016/j.jembe.2020.151489

https://doi.org/10.1016/j.jembe.2020.151489

http://crossmark.crossref.org/dialog/?doi=10.1016/j.jembe.2020.151489&domain=pdf

Journal of Experimental Marine Biology and Ecology 535 (2021) 151489

2

Porzio et al. 2011). Discrepancies in the literature may depend on
species-specific calcification mechanisms, photophysiology, location of
calcification site, and calcium carbonate (CaCO3) crystal form and
mineral content (reviewed in Basso 2012; Hofmann and Bischof 2014;
Koch et al. 2013).

In marine macroalgae, CaCO3 precipitation occurs in areas that are
isolated or semi-isolated from bulk seawater where the saturation state
(ΩCaCO3) can be elevated to promote calcification (Borowitzka and
Larkum 1987). Calcification typically occurs in the cell walls of Rho-
dophyta (red macroalgae), such as crustose coralline algae (CCA), and
other Rhodophyta families (e.g., Peyssonneliacea) (Adey et al. 2013;
Basso 2012). In the Chlorophyta (green macroalgae), calcification oc-
curs in sheaths surounding filaments or specific compartments con-
nected to external seawater by diffusive channels (Borowitzka and
Larkum 1987, 1976). CCA are thought to be the most sensitive to
declining pH and elevated pCO2 due to the proximity of their calcifying
sites to overlying bulk seawater and the high magnesium-calcite content
of their crystals. Magnesium concentrations in CaCO3 lattice have a
positive relationship with temperature and higher concentrations of Mg
result in a relatively more soluble polymorph of CaCO3 (Kamenos et al.
2009; Kamenos and Law 2010; Mccoy and Kamenos 2015). A number of
studies have shown negative effects of elevated pCO2 and low pH on
CCA calcification (Anthony et al. 2008; Basso 2012; Comeau et al. 2019;
Diaz-Pulido et al. 2014; Gao et al. 1993; Kato et al. 2014; Noisette et al.
2013); however, other studies imply CCA resistance (Comeau et al.
2018, 2017, 2013; Cornwall et al. 2017; Dutra et al. 2015; Ries et al.
2009). Chlorophytes, and some rhodophytes, have an aragonite poly-
morph of CaCO3, which is less soluble than the high magnesium poly-
morph found in CCA (Borowitzka and Larkum 1987). Precipitating a less
soluble CaCO3 polymorph within semi-isolated compartments may be
advantageous to resist elevated pCO2 and low pH (Comeau et al. 2013;
Peach et al. 2017b, 2016; Ries 2011; Vogel et al. 2015a), yet there are a
number of studies showing lower net calcification under declining pH
and elevated pCO2 conditions (Meyer et al. 2016; Price et al. 2011). In
addition to the potential effects of morphology and polymorphs,
photosynthesis affects calcification in marine macroalgae (Diaz-Pulido
et al. 2007; Hofmann and Bischof 2014; Koch et al. 2013; Porzio et al.
2011; Raven and Hurd 2012).

Photosynthesis has been shown to increase calcification in marine
macroalgae (Borowitzka and Larkum 1987; De Beer and Larkum 2001;
Gao et al. 1993; Koch et al. 2013; Pentecost 1978; Semesi et al. 2009;
Wizemann et al. 2014), but the influence of increasing pCO2 and [H

+] on
the coupling of these two processes is only recently being disentangled
(Brown et al. 2019; Comeau et al. 2018; Hofmann et al. 2016; McNicholl
et al. 2020, 2019). A majority of marine macroalgae use carbon
concentrating mechanisms (CCMs) to saturate RuBisCO with CO2 for
photosynthesis (Raven and Hurd 2012). In the process of HCO

3

− uptake,
HCO3

− dehydrogenation to CO2 and OH
− , catalyzed by external carbonic

anhydrase (CAext), can neutralize H
+ and raise the macroalgal surface

pH. Immediate (seconds) light-triggered pH increase in macro-and
micro-algal surfaces detected with microsensors, combined with
photosynthetic inhibitors, provides evidence that photosynthesis and
light are major drivers of pH control at the seawater-cell surface inter-
face (Chrachri et al. 2018; Cornwall et al. 2015, 2013; De Beer and
Larkum 2001; Hofmann et al. 2016; McNicholl et al. 2019). Presence
and maintenance of a high thalli surface pH may support calcification
under declining pH and elevated pCO2 (Cornwall et al. 2014; Hofmann
et al. 2016; McNicholl et al. 2019). In addition to photosynthesis, light-
triggered H+ transport pumps independent of photosystem II (PSII) have
been identified in several species of macroalgae and may facilitate
calcification (De Beer and Larkum 2001; Hofmann et al. 2016; McNi-
choll et al. 2019). Electron microscopy of epithallial cells show in-
vaginations that have been postulated to promote proton pumping in
coralline algae during decalcification/recalcification to support thalli
growth (Pueschel et al., 2005). Thus, photosynthesis and active proton
pumping will likely play an important role for continued calcification

(or dissolution) processes.
While calcification continues to occur in some marine macroalgae in

the dark, rates are typically reduced or become net negative (Chisholm
2000; El Haïkali et al. 2004; Vogel et al. 2015b), and may rely on
accumulated energy stored during periods of irradiance (Mccoy and
Kamenos 2015). Dark dissolution is likely driven by lower pH at calci-
fication sites (Borowitzka and Larkum 1976; Wizemann et al. 2014).
Amplification of this effect may occur with lower seawater carbonate ion
concentrations [CO3

2− ] and lower ΩCaCO3, as well as increased seawater
[H+] and pCO2 in the bulk seawater (Comeau et al. 2012; McNicholl
et al. 2020). A buildup of external CO2 or H

+ in the dark may prevent
removing H+ from the calcifying space against a higher [H+] concen-
tration in seawater (Cyronak et al. 2015; Jokiel 2011), although some
biotic control and dissolution itself can buffer pH of the bulk seawater at
the surface in the dark. Internal cellular acid-base regulation may also
become difficult in the dark when the electrochemical gradient proton
motive force reverses from passive diffusive efflux of H+ (pHsw:pHcell
8.2:7.2) to requiring active H+ transport (pHsw:pHcell 7.8:7.2), as
exemplified in calcifying phytoplankton with proton channels (Taylor
et al. 2012). Higher energetic demands for proton pumps or lack of H+

regulation could shift net calcification to net dissolution at night even
for those groups, such as the CCA, with a known potential to biotically
control calcification/decalcification as part of their life history (Pueschel
et al., 2005).

The objective of this study was to determine the effects of a 0.4 pH
decline (pH 7.7) from the current ambient pH (8.1), as predicted for
2100 (Gehlen et al. 2014; Hartin et al. 2016), on net calcification rates of
nine tropical reef macroalgae. Species examined were characterized by
diverse calcification locations, CaCO3 content, polymorph, CCMs, and
site-specific irradiance levels. This included five species from a high-
light patch reef on the Florida Keys Reef Tract (FKRT) and four spe-
cies from under reef wall ledges with low light on Little Cayman Island
(LCI). Calcification rates independent of photosystem-II (PSII) were also
investigated for FKRT species. All calcification experiments were con-
ducted in the light and dark. We hypothesized that algae from high-light
patch reefs would continue to calcify at elevated pCO2 due to their high
potential to raise surface pH through photosynthesis (McNicholl et al.
2019) and calcifying space pH (Cornwall et al. 2017). Low-light species
were predicted to have low photosynthetic rates; therefore, calcification
rates would be reduced at elevated pCO2. Further, we postulated that
elevated pCO2 would not lower net calcification rates in FKRT species
shown to exhibit light-triggered thalli H+ regulation independent of PSII
based on microsensor studies (McNicholl et al. 2019). We proposed that
net calcification rates would be lower in the dark and this decline would
be amplified at low pH.

2. Materials & methods

2.1. Calcified macroalgae collection and site parameters

2.1.1. Florida keys reef tract (high light environment)
Intact individuals of five dominant calcifying algae, including 3

Rhodophyta (Neogoniolithon strictum, Jania adhaerens, CCA) and 2
Chlorophyta (Halimeda scabra, Udotea luna) were collected (between
March – June 2018) from a shallow Looe Key patch reef (~3–5 m)
(24.62055◦ N, 81.37078◦ W) on the FKRT (Fig. 1a, Fig. 2a-e). Neo-
goniolithon strictum and J. adhaerens were collected by hand from the
benthos, while H. scabra and U. luna were collected by inserting a dive
knife into the surrounding carbonate sediment and lifting out the thalli
with rhizoids intact. CCA-covered carbonate rubble (Fig. 2c) was also
collected.

Field light levels were measured during collections at midday
(~11:00–13:00) just above the benthos with a 4π spherical PAR quan-
tum sensor (LI-193, LI-COR Inc.). The shallow patch reef site had high-
irradiance (~800 μmol photo m− 2 s− 1) based on average midday mea-
surements during collections. Experiments with high-light species from

C. McNicholl and M.S. Koch

Journal of Experimental Marine Biology and Ecology 535 (2021) 151489
3

the FKRT were conducted under a full-spectrum LED light (Kessil,
A360W E-Series Tuna Sun) at 500 μmol photons m− 2 s− 1 (LI-COR
Quantum 4 π Light Sensor). While this was lower than mid-day field
irradiance at the benthos, 500 μmol photons m− 2 s− 1 PAR approximated
light levels that saturated photosynthesis in these species across a range
of pH (Zweng et al. 2018).

Seawater in situ chemistry was measured or calculated for each
macroalgal collection (n = 3) (Table 1a). Site pH (Orion A211,
8302BNUMD; calibrated with NBS standards, Thermo Fisher Scienti-
fic®), conductivity (salinity = 35.5) and temperature (28.5 ± 1.1 ◦C)
(YSI 650 MDS) were determined in the field. Water samples (n = 3;
60 mL) were collected and total alkalinity determined within 48 h or
fixed with HgCl2 (0.02%) and measured within eight weeks. For clarity
and continuity, measurements for pH are reported on the NBS scale since
correction to total scale using the TRIS buffer was not available for ex-
periments conducted on LCI. Total alkalinity (TA) was measured by
open-cell titration (Metrohm Titrando® 888) with 0.01 N HCl. Certified
reference material (CRM Batch #156; Dickson Lab, Scripps Institute of

Oceanography) was also run for each batch of TA samples. The CRM
offset was used to correct TA readings from each batch. This way the
data had no systematic bias from the true value. A certified standard
TRIS buffer was used to calibrate pH for total alkalinity analysis (TRIS
buffer, Dickson Lab, Scripps Institute of Oceanography: pH = 8.21,
mV = − 64.9, 25 ◦C and 35 salinity). TA measurements were performed
in triplicate unless the initial two measurements were within
±5 μmol kg− 1 of each other. Total alkalinity, temperature, conductivity,
and pHNBS data were used to calculate DIC speciation (CO2SYS, Pierrot
et al. 2006 with; K1, K2 from Mehrbach et al., 1973 refit by Dickson and
Millero, 1987) for each experiment.

Following collections, algal samples were immediately transported
(4 h) in aerated coolers to Florida Atlantic University (FAU) in Boca
Raton, FL. Samples were separated by species and acclimated for 48 h in
9 L aquaria held within a large mesocosm tank (~500 L) over which 2
(100 cm length x 10 cm width) full-irradiance spectrum fixtures
(BuildMyLED Inc.) were hung. To provide all 9 aquaria with a similar
light field, the fixtures were hung ~0.5 m above the tank and provided

Fig. 1. Collection sites at (a, top panel) a shallow (3–5 m) high irradiance patch reef of the Florida Keys Reef Tract (Looe Key Reef; 24.62055◦ N, 81.37078◦ W) and
(b, middle panel) a deeper (20 m) low irradiance wall reef on Little Cayman Island (Rockbottom Reef; 19.7025◦N, 80.05694◦ W). The (c) experimental setup for total
alkalinity anomaly incubations where macroalgae were raised above the bottom of the beakers on a perforated disk (see insert) to allow for continued, slow stirring
below the macroalga. Irradiance on the reefs was measured using a 4 π sensor; on the deep wall reefs (d) an underwater data logger was deployed.

C. McNicholl and M.S. Koch

Journal of Experimental Marine Biology and Ecology 535 (2021) 151489

4

Fig. 2. Species from the (a-e) shallow (3–5 m) high
irradiance patch reef of the Florida Keys Reef Tract and
(f-i) deeper (20 m) low irradiance wall reef on Little
Cayman Island used to examine the effects of elevated
pCO2 and low pH predicted for 2100 on net calcifica-
tion rates in light and dark experiments. Species from
the FKRT (a) Neogoniolithon strictum, (b) Jania adhae-
rens, (c) CCA, (d) Halimeda scabra, (e) Udotea luna and
LCI Peyssonneliaceae (f) Peypink, (g) Peyred, (h) Hal-
imeda copiosa and (i) Halimeda goreauii.

C. McNicholl and M.S. Koch

Journal of Experimental Marine Biology and Ecology 535 (2021) 151489

5

~250 μmol photons m− 2 s− 1 to each aquarium on a 12:12 light:dark
cycle. All aquaria were semi-immersed (~80%) in the mesocosm tank
with water maintained at 29 ◦C, the average temperature for the FKRT
during summer (Kuffner et al. 2015). Water in the aquaria was aerated
and circulated via small submersible pumps and replenished (75%)
every 2 d with seawater from the flow-through seawater system at the
FAU marine lab (coastal Atlantic Ocean). Experiments were completed
within 5 d of collections.

2.1.2. Little Cayman Island reef (low light environment)
Calcified macroalgae were collected from a low-light reef wall

(Fig. 1b) on LCI and experiments conducted at the Little Cayman
Research Center (LCRC) (July 2018). Four species, including 2 Rhodo-
phyta prostrate lobed crustose algae (Family Peyssonneliaceae; Peyred
and Peypink) and 2 Chlorophyta (Halimeda goreauii and Halimeda copiosa)
were collected under ledges and crevices along the upper reaches of a
reef wall (Rockbottom ~20 m; 19.7025◦N, 80.05694◦ W; Fig. 2 f-i). The
Halimeda species hung under ledges (Fig. 1b, Fig. 2h-i) and were
collected from the holdfast removing the loosely attached filaments.
Peyssonneliaceae samples (Fig. 2 f-g) were carefully removed with a
small chisel (20 cm) where the crustose lobes were attached to the
substrate. While macroalgae were being collected, light was measured
(LI-COR Quantum 4 π Light Sensor) in situ under ledges (Fig. 1d), col-
lecting data during midday (~11:00–13:00) for ~5 min with s− 1 interval
(RBRsolo3 single channel data logger). The range of light under ledges at
the collection site was ~5 to 50 μmol photons m− 2 s− 1. Macroalgae were
kept in aerated seawater from the reef at low light and immediately
transportation to LCRC (< 2 h). At LCRC, algae were separated into 4 L aquaria, kept aerated in the shade (~ 50 μmol photons m− 2 s− 1) with natural sunlight and experiments run within 24 h. Seawater was collected and immediately analyzed for pH and salinity upon returning to the lab, and samples were fixed with HgCl2 (0.02%) for total alkalinity analysis within 8 weeks.

2.2. Elevated pCO2 experiments

Net calcification rates were determined for macroalgae in seawater

adjusted to pH (7.7) predicted for 2100 (Gehlen et al. 2014; Hartin et al.
2016) and ambient controls (8.1). Final pH and pCO2 treatment levels
attained (see results) were determined from the initial and end mea-
surement of each light (between 0800 and 1500) and dark (between
1900 and 0300) experiment. Calcification rates of FKRT and LCI species
were determined using the TAA technique. The TAA technique for
determining calcification rates in marine calcifiers is based on the
changes in seawater TA over time. The TAA target was 3–10 times the
accuracy of the method (~10 μmol kg− 1) and within 10% of TA (Lang-
don et al. 2010). Based on this protocol, experimental incubation time
for high-light species from the FKRT (n = 3–5) was 1–3 h in the light,
depending on species, and 4 h in the dark. Incubation time for low-light
species from LCI (n = 4) was 4.5 h in the light and 5.5 h in the dark.
Blank runs with seawater were also conducted. Experiments were con-
ducted in glass beakers (150–250 mL) covered with parafilm and
secured with rubber bands to reduce atmospheric gas exchange (Fig. 1c)
according to Chisholm and Gattuso (1991). Individual thalli were sus-
pended approximately 2 cm above the bottom of the beaker on a
perforated disk with a stirbar underneath and flow created with a stir-
plate (Fig. 1c insert).

Experimental seawater was filtered (0.45 μM) and brought to tem-
perature (29 ◦C) in a waterbath before assigning treatment. Low pH
treatment was obtained by bubbling seawater with pure CO2 prior to
incubations. Initial and post-incubation pH, O2 (Orion A329), and
temperature (Orion A211, 8302BNUMD) were recorded. Algae were
acclimated to experimental seawater for 15 min. A relatively short
acclimation time was needed to keep carbonate chemistry and △pH to a
minimum before incubations. Low-light experiments with LCI species
were conducted using the same experimental setup and protocols with
the exception that these species were incubated at 50 μmol photons
m− 2 s− 1.

Ambient and low pH treatment runs were randomized so ~50% of
the runs were low pH treatments, followed by ambient controls, and the
inverse for the other 50%. In this experiment, repeated measures were
used (i.e., treatments were applied sequentially to the same individual
thalli). When similar TAA experiments were run during 45Ca experi-
ments (McNicholl et al. 2020), similar results were found for TAA in

Table 1
Summary of (a) average seawater carbonate chemistry during incubations averaged across experiments in light/dark and ambient (A) or low (L) pH. In situ seawater
conditions (n = 3) from collection sites Florida Keys Reef Tract (FKRT) and Little Cayman Island (LCI) shown. Carbonate chemistry parameters were calculated using
CO2SYS (Pierrot et al. 2006) applying experimental seawater temperature (29 ◦C) and salinity (35.5), and (b) change (△) in seawater pH, total alkalinity (TA) and
oxygen (O2) during incubations in ambient and low pH experiments in the light and dark averaged across all experiments (n = 38). Means ± SD. Details of carbonate
chemistry and △pH, △TA and △O2 are presented for all experiments by species and treatments in supplemental tables (Table S1, S2).
Summary of (a) average seawater carbonate chemistry during incubations averaged across experiments in light/dark and ambient (A) or low (L) pH. In situ seawater
conditions (n = 3) from collection sites on the Florida Keys Reef Tract (FKRT) and Little Cayman Island (LCI) reefs are shown. Carbonate chemistry parameters were
calculated using CO2SYS (Pierrot et al. 2006) applying experimental seawater temperature (29 ◦C) and salinity (35.5).The (b) change (△) in seawater pH, total
alkalinity (TA) and oxygen (O2) in the light and dark are averaged across all experiments (n = 38). Means ± SD. Details of carbonate chemistry and △pH, △TA and
△O2 are presented for all experiments by species and treatments in supplemental tables (Table S1, S2).

(a) pH pCO2
(μatm)

HCO3
−

(μmol kg− 1)
CO3

2−

(μmol kg− 1)
TA

(μmol kg− 1)
ΩCa ΩArag

Experimental
Light – A

Light – L
8.07 ± 0.06
7.71 ± 0.05

367 ± 64
992 ± 140

1665 ± 76
1981 ± 58

248 ± 23
130 ± 12

2283 ± 29
2302 ± 39

6.0 ± 0.6
3.1 ± 0.3

4.0 ± 0.4
2.1 ± 0.2

Dark – A
Dark – L

8.00 ± 0.03
7.62 ± 0.03

451 ± 34
1239 ± 112

1760 ± 31
2062 ± 29

222 ± 11
111 ± 8

2305 ± 31
2329 ± 31

5.4 ± 0.3
2.7 ± 0.2

3.6 ± 0.2
1.8 ± 0.1

In Situ
FKRT

LCI
8.10 ± 0.02
8.05 ± 0.03

343 ± 11
402 ± 37

1662 ± 17
1752 ± 39

259 ± 4
243 ± 13

2292 ± 31
2348 ± 24

6.2 ± 0.1
5.9 ± 0.3

4.2 ± 0.1
3.9 ± 0.2

(b) Ambient pH Experiments Low pH Experiments

△ pH
△ TA

(μmol kg− 1)
△ O2

(mg L− 1)
△ pH

△ TA
(μmol kg− 1)

△ O2
(mg L− 1)

Experimental
Light <0.01 ± 0.09 − 70 ± 41 1.60 ± 0.93 0.09 ± 0.09 − 43 ± 50 1.82 ± 0.90 Dark − 0.11 ± 0.03 − 17 ± 12 − 0.42 ± 0.33 − 0.08 ± 0.03 22 ± 16 − 0.51 ± 0.31

C. McNicholl and M.S. Koch

Journal of Experimental Marine Biology and Ecology 535 (2021) 151489

6

FKRT species using completely independent thalli for each run.

2.3. PSII inhibitor experiments

To examine light-dark PSII-independent calcification rates, an in-
hibitor, 3-(3,4-Dichlorophenyl)-1,1-dimethylurea (DCMU), was added
to the seawater according to Hofmann et al. (2016), after De Beer and
Larkum (2001) and Borowitzka and Larkum (1976). DCMU was also
highly effective at arresting photosynthesis in microsensor studies,
eliminating O2 flux from the diffusive boundary layer (100 μm) at the
thalli surface (McNicholl et al. 2019). A stock solution of 0.05 M DCMU
was added to reach a concentration of 4 μM. Algae were left in DCMU-
amended seawater until no change in O2 production was detected
upon illumination (~30 min). Incubations were conducted (as above) at
both pH levels and light/dark conditions with DCMU. DCMU experi-
ments were only performed on species from the FKRT species due to time
constraints on LCI.

2.4. Calculation of net calcification and chemistry analysis

Net calcification (Gnet) rates in μmol CaCO3 g dwt− 1 h− 1 were
calculated from TA changes (△TA) during the experiments based on the
following equation (Eq. 1):

Gnet = − 0.5ρw
∆TA*v
Wa*t

(1)

where: ρw is seawater density (kg L− 1) and ∆TA (μmol kg− 1) is the final
TA minus the initial TA, v is the volume of seawater (L), Wa is the algal
dry weight (g) and t is the incubation time in hours. Calcification rates
were normalized to dry weight (60 ◦C) for FKRT species with the
exception of CCA that had a complex 3-D structure. CCA was normalized
to g of a flexible 2-D surface (foil) according to Marsh (1970). Calcifi-
cation rates of Halimeda spp. from LCI were normalized to dry weight as
above and to cm− 2 of thalli surface for LCI Peyssonneliaceae. Daily
calcification rates were calculated by combining light (LGnet) and dark
(DGnet) net calcification rates using a 12:12 light:dark cycle, as a first-
order estimate of rates over 24 h (Eq. 2).

Daily Gnet = (LGnet*12) + (DGnet*12) (2)

2.5. Photosynthesis and respiration

Photosynthesis and respiration rates were determined from initial
and final dissolved O2 measurements (optical probe) using data from
incubation runs in the light and dark, respectively. The O2 flux rates
were normalized to dry mass (g) or surface area, as described above, and
adjusted to seawater volume and time.

2.6. Statistical analysis

A repeated measures two-way ANOVA was performed (SigmaPlot
v13.0, Systat Software Inc.) to compare calcification rates across light/
dark at ambient and low pH treatments for experiments with and
without DCMU. The assumptions of normality and homogeneity of
variance were tested using the Shapiro-Wilkes and Brown-Forsythe tests,
respectively. The assumption of sphericity in repeated measures was
tested using the Mauchly’s test. Differences amongst means were
established using the Holm-Sidak post-hoc test. The effect of pH treat-
ments on the calculated daily calcification rates were determined using a
t-test. Regression analysis was used to establish the relationship between
net photosynthesis and calcification (SigmaPlot v13.0, Systat Software
Inc.). Significance levels were established at p < 0.05 unless otherwise stated.

3. Results

3.1. Carbonate chemistry and △pH, △TA and △O2

Carbonate chemistry, treatment and in situ pH and pCO2, △pH,
TA and △O2 are summarized in Table 1. All data by species and

treatments are presented in supplemental tables (Tables S1, S2). The
resulting average pH and pCO2 across experiments and treatments was
7.67 and 1116 μatm, respectively (Table 1). The average ambient pH
and pCO2 for controls were 8.04 and 409 μatm, respectively (Table 1).
The averages of the initial and end pCO2 during all experiments was
approximately 3-fold higher in the elevated pCO2 treatments compared
to controls, resulting in a pH of 7.71 for the low pH treatment compared
to 8.07 for ambient pH (Table 1a). The dark experiments were ~ 0.08 pH
lower than in the light due to metabolic differences, but the values were
within the pH variance found in the light experiments (Table 1a). The
concentrations of CO3

2− and saturation state of CaCO3 declined ~50% in
the high pCO2 treatments but remained above CaCO3 saturation (Ω > 1)
for all experiments (Table 1a). The △pH and △O2 across experiments
indicate only modest changes in chemistry occurred throughout the
experiments (Table 1b). The △pH was on average ± 0.03 to ±0.09
showing that our treatments were close to those applied at the initiation
of the experiment. This is also indicated by a small change in metabolic
O2, which was similar in the light for both high and low pH treatments
(Table 1b). Average TA changes in the light were at least 3 times our
accuracy for CRM TA (± 10 μmol kg− 1) and remained significantly lower
than 10% of seawater TA. Further, the TA data for all the experiments
highlight our general result that in the light △ TA was negative (indi-
cating net calcification occurred) (Table 1b).

3.2. High light species (FKRT) calcification

Greater net calcification rates were found in the light compared to
the dark for all species (Fig. 3) based on 2-way ANOVA main effects. Net
calcification rates were similar at low and ambient pH for N. strictum
(Holm-Sidak, t = 1.38 p = 0.214) and J. adhaerens (Holm-Sidak, t = 1.99
p = 0.085) in the light (Fig. 3a,b). Halimeda scabra net calcification rates
increased 16% at low pH in the light (Fig. 3d; Holm-Sidak, t = 2.54,
p = 0.035), indicating a positive response to elevated pCO2. In contrast,
U. luna (Holm-Sidak, t = 3.07 p = 0.018) and CCA (Holm-Sidak, t = 4.56
p = 0.028) net calcification rates significantly declined at low pH rela-
tive to controls in the light with CCA eliciting the strongest negative
effect (Fig. 3c,e).

Only N. strictum and J. adhaerens showed significant positive net
calcification rates in the dark and only at ambient pH (Fig. 3a,b). Dark
calcification rates for these two species were only a tenth the rates
observed in the light. Even at ambient pH, U. luna exhibited net disso-
lution in the dark (Fig. 3c), while H. scabra and CCA had net calcification
rates approximating zero (Fig. 3d,e). Neogoniolithon strictum,
J. adhaerens, and H. scabra (Holm-Sidak, t = 2.73 p = 0.032, t = 2.34
p = 0.050, t = 2.71 p = 0.027, respectively) had significantly lower net
calcification rates in the dark at low pH compared to ambient pH
(Fig. 3a,b,d). CCA had 87% less net calcification on average at low pH
compared to ambient pH in the dark, although not significant (Holm-
Sidak, t = 1.59 p = 0.226).

The two species that were negatively affected by low pH in the light,
U. luna and CCA, had lower daily calcification rates at low pH relative to
ambient controls (Fig. 3c,e; t-test, t = 2.65 p = 0.038, t = 5.27 p = 0.006,
respectively). In contrast, the three species with no significant negative
low pH effect on net calcification in the light, N. strictum (t-test, t = 1.21
p = 0.262), J. adhaerens (t-test, t = 1.16 p = 0.278), and H. scabra (t-test,
t = 0.03 p = 0.977), had similar daily calcification rates in low and
ambient pH treatments (Fig. 3a,b,d). Halimeda scabra’s significantly
lower net calcification rate in the dark at low pH was offset by increased
calcification rates in the light at low pH, resulting in no difference in
daily calcification rates (Fig. 3d).

C. McNicholl and M.S. Koch

Journal of Experimental Marine Biology and Ecology 535 (2021) 151489

7

3.2.1. Photosynthesis and calcification
Net photosynthesis and respiration were not different between the

low and ambient pH for any species from the FKRT or LCI (Table 2).
Species from the FKRT that exhibited a broad range in photosynthetic
rates amongst individuals exhibited a strong correlation between net
photosynthesis and net calcification rates (Fig. 4). Linear relationships
between net photosynthesis and calcification rates for N. strictum and
H. scabra were stronger at low pH (R2 = 0.96; R2 = 0.94) compared to
ambient pH (R2 = 0.68; R2 = 0.64), respectively (Fig. 4a,c), while for
J. adhaerens the relationships were relatively similar (R2 = 0.87,
R2 = 0.98; Fig. 4b). The other two species did not have a broad range in
photosynthetic and calcification rates, thus no relationship could be
established.

3.2.2. Inhibitor experiments (FKRT)
The photosynthesis inhibitor (DCMU) arrested O2 flux in the light for

all species, providing confidence that calcification in the presence of
DCMU was non-PSII light-dependent calcification (Table 2). Calcifica-
tion was stimulated in the light without PSII in three FKRT species,

including N. strictum, J. adhaerens, and H. scabra (Fig. 5a,b,d). These
species maintained a relatively high percentage (22 to 34%) of the
calcification rates attained in the light without PSII inhibition (Fig. 3a,b,
d). This light-triggered non-PSII net calcification was significantly
greater than calcification rates measured in the dark at ambient pH
(Holm-Sidak, t = 3.58 p = 0.012, t = 3.98 p = 0.004, t = 3.00 p = 0.032,
N. strictum, J. adhaerens, and H. scabra, respectively). However, non-PSII
light-dependent calcification rates were not sustained at low pH.
Calcification rates at low pH without PSII were similar in the light and
dark (Holm-Sidak, t = 0.43 p = 0.681, t = 1.84 p = 0.104, t = 1.01
p = 0.351, N. strictum, J. adhaerens, and H. scabra, respectively). No
significant light-triggered calcification with DCMU was observed for
U. luna (Two-way ANOVA, F1,4 = 1.55 p = 0.281) or CCA (Two-way
ANOVA, F1,4 = 0.065 p = 0.812). For all species examined, no net posi-
tive calcification rates were measured in the presence of DCMU at low
pH in the dark (Fig. 5).

Fig. 3. Net calcification rates (n = 3–5) of Florida Keys Reef Tract macroalgal species (a = Neogoniolithon strictum, b = Jania adhaerens, c = Udotea luna, d = Halimeda
scabra, e = CCA) in the light (500 μmol photons m− 2 s− 1) and dark at ambient and low pH. Totals are daily net calcification rates calculated by combining light and
dark calcification rates at the respective treatment pH and normalized to 24 h. Means ± standard errors are shown. Different lowercase letters represent significant
differences from a two-way ANOVA with Light x pCO2 treatments and a post-hoc Holm-Sidak to compare between means (P < 0.05). *CCA calcification rates are normalized to g of flexible 2-D surface due to high 3-D complexity. Differences in daily calcification rates and CCA within light treatments were determined by t-tests; daily differences shown with capital letters (P < 0.05).

C. McNicholl and M.S. Koch

Journal of Experimental Marine Biology and Ecology 535 (2021) 151489

8

3.3. Low light species (LCI) calcification

Although the low-irradiance Halimeda species from LCI reefs were
incubated at 10-fold lower irradiance and had 10 times lower photo-
synthetic rates compared to FKRT Halimeda species (Table 2), their
calcification rates in the light (Fig. 6a,b) were similar (Fig. 3d). At
ambient pH, net calcification rates were significantly higher in the light
relative to the dark (Fig. 6; Holm-Sidak, t = 4.92 p = 0.003, t = 3.28
p = 0.038, t = 3.37 p = 0.023, H. goreauii, H. copiosa, Peyr, respectively)
with the exception of Peyp, that approached significance (Fig. 6d; Holm-
Sidak, t = 2.59 p = 0.073). Lower dark calcification rates compared to
the light were consistent with results from the FKRT species (Fig. 3).
However, in contrast to Halimeda from the FKRT, relatively high positive
net calcification rates were maintained in the dark by Halimeda species
from LCI reefs, H. goreauii (39%) and H. copiosa (22%) (Fig. 6a,b). These
dark calcification rates are more than twice those of N. strictum and J.
adhaerens from the FKRT (Fig. 3a,b). Calcification rates at low pH in the
light decreased by 49% and 28% for H. goreauii and H. copiosa, respec-
tively (Fig. 6a,b), but was only significant for H. goreauii (Holm-Sidak,
t = 3.19 p = 0.021), likely due to high variance in H. copiosa (Holm-
Sidak, t = 1.78 p = 0.149). Although both H. goreauii and H. copiosa had
on average ~ 50% lower daily calcification rates at low pH in the light,
the differences only approached significance for H. goreauii (t-test,
t = 2.15 p = 0.0748) and was not significant for H. copiosa (t-test,
t = 1.18 p = 0.281). Peyred from LCI only exhibited positive net calcifi-
cation rates in the light at ambient pH and showed net dissolution at low
pH in the light and dark (Fig. 6c). Although similar trends were
observed, low pH only significantly reduced Peypink (Holm-Sidak,
t = 3.17 p = 0.047) net calcification rates in the dark, but not Peyred
(Holm-Sidak, t = 0.800 p = 0.460). Daily calcification rates were

significantly lower at low compared ambient pH for Peypink (t-test,
t = 2.93 p = 0.026) and approached significance for Peyred (t-test,
t = 2.30 p = 0.061) (Fig. 6c,d).

4. Discussion

The most consistent negative effect of low pH and elevated pCO2 on
calcification rates in tropical macroalgae examined from FKRT and LCI
occurred in the dark, albeit effects in the light controlled daily calcifi-
cation rates. Most of the species examined (89%) had calcification rates
of zero or net dissolution in the dark under 2100 predictions for pH and
ocean carbonate chemistry. Other experimental and field studies also
indicate negative effects of elevated pCO2 and low pH on dark calcifi-
cation rates. Dark calcification rates of Halimeda opuntia decreased
167% at a low pH (~7.8) tropical CO2 seep site compared to non-seep
adjacent control areas (Vogel et al. 2015a). These data are comparable
to the 171% decrease in dark calcification rates observed at low pH for
the three Halimeda species in this study. A temperate coralline alga
(Lithothamnion glaciale) with net positive calcification rates in the dark at
ambient pH, also exhibited net dissolution when exposed to low pH in

Table 2
Net photosynthesis and respiration rates during the light and dark calcification
incubations at ambient and low pH and with and without a photosystem II in-
hibitor (DCMU). All data are normalized to gram dry weight per hour with the
exception of CCA from the Florida Reef Tract and Peyssonneliaceae from Little
Cayman Island (see below). Mean +/− SE (n = 5–3).

Oxygen (μmol O2 g dwt− 1 or cm− 2 h− 1)*

Net Photosynthesis Respiration

Amb pH Low pH Amb pH Low pH

Florida Reef
Tract (No
DCMU)

N. strictum 8.66 ± 1.66 8.83 ± 1.42 -0.76 ± 0.08 -0.96 ± 0.17
J. adhaerens 15.47 ± 4.13 19.29 ± 4.09 -2.39 ± 0.49 -2.97 ± 0.70
U. luna 12.85 ± 1.51 12.54 ± 0.95 -2.08 ± 0.34 -2.10 ± 0.48
H. scabra 8.12 ± 1.95 8.40 ± 2.12 -1.30 ± 0.22 -1.36 ± 0.45
*CCA 37.24 ± 13.08 48.30 ± 13.59 -6.28 ± 1.84 -8.62 ± 2.79

Florida Reef
Tract (+
DCMU)

N. strictum -1.09 ± 0.42 -1.75 ± 0.44 -1.29 ± 0.21 -1.37 ± 0.26
J. adhaerens -2.63 ± 1.08 -3.81 ± 0.76 -2.92 ± 0.50 -4.13 ± 0.70
U. luna -2.70 ± 0.88 -3.56 ± 0.24 -3.31 ± 0.71 -3.39 ± 0.99
H. scabra -2.51 ± 0.73 -2.94 ± 0.52 -2.45 ± 0.71 -2.31 ± 0.53
*CCA -11.76 ± 2.64 0.02 ± 0.35 -14.40 ± 4.62 -11.19 ± 1.27

Little
Cayman
Island (No
DCMU)

H. goreauii 1.37 ± 0.98 2.55 ± 0.84 -0.49 ± 0.19 -0.33 ± 0.31
H. copiosa 3.34 ± 0.94 3.26 ± 1.00 -0.53 ± 0.32 -0.92 ± 0.32
*Peyred 0.19 ± 0.01 0.19 ± 0.04 -0.05 ± 0.01 -0.05 ± 0.03
*Peypink 0.12 ± 0.03 0.17 ± 0.03 -0.01 ± 0.01 -0.06 ± 0.02

* CCA from the Florida Reef Tract normalized to g of flexible 2-D surface due
to high 3-D complexity and Peyssonneliaceae; Peyred and Peypink from Little
Cayman Island to cm− 2 of thalli surface.

Fig. 4. The linear relationship between net photosynthetic and calcification
rates (n = 5) in three Florida Reef Tract species (a = Neogoniolithon strictum,
b = Jania adhaerens, c = Halimeda scabra). Equations and R2 for linear re-
gressions are shown for ambient pH (circles with solid line) and low pH (tri-
angles with dashed line) treatments.`

C. McNicholl and M.S. Koch

Journal of Experimental Marine Biology and Ecology 535 (2021) 151489
9

the dark (Kamenos et al. 2013). Reduced net calcification by 164% in
macroalgae (H. opuntia) from the Great Barrier Reef at low pH in the
dark was interpreted as a negative amplifying effect of low pH (Vogel
et al. 2015b). The impact of low pH and elevated pCO2 on preferentially
nighttime net calcification (Kamenos et al. 2013; Venn et al. 2019; Vogel
et al. 2015b; this study) necessitates a greater understanding of these
mechanisms.

The constraints on net calcification rates in the dark at low pH is
likely attributable to greater rates of dissolution. This conjecture is
reasonable given McNicholl et al. (2020) found in 45Ca experiments
either no significant difference between gross calcification rates in the
dark between low (7.7) and ambient (8.1) pH, or a shift from positive
gross calcification rates to net dissolution at low pH in the dark, in a
majority of the FKRT species examined herein (N. strictum, J. adhaerens,
H. scabra, U. luna). This was the case, even though net calcification rates
significantly declined at low pH in the dark in all four of the FKRT
species examined (McNicholl et al. 2020). McNicholl et al. (2020) also
established a strong relationship (R2 = 0.82) between increasing TA and

loss of 45Ca from pre-labelled 45CaCO3 thalli only in the low pH treat-
ment, suggesting dissolution. These data indicate that the ability to form
new CaCO3 is not the primary factor constraining net calcification rates
in the dark at low pH for the majority of FKRT species examined herein.
Further, respiration rates cannot account for the increased dissolution at
low pH in darkness. None of the species examined exhibited greater
respiration rates in the dark at low pH. Comeau et al. (2016) also found
respiration rates to be insensitive to elevated pCO2 in 6 coral and 6
macroalgal species from reefs in Moorea (French Polynesia). Other
macroalgal studies support the conclusion that respiration rates do not
increase in response to low pH conditions in the dark (Kamenos et al.
2013; Martin et al. 2013a; Semesi et al. 2009; Zou et al. 2011; Zweng
et al. 2018). A general amplifying effect of low pH on dissolution in the
dark was observed in both high- and low-light macroalgae in the present
study, regardless of species-specific photophysiology, calcification
location, taxonomy, or thalli CaCO3 polymorph, suggesting a funda-
mental relationship that necessitates further research.

In contrast to results in the dark, low pH had no significant negative

Fig. 5. Net calcification rates (n = 3–5) of Florida Reef Tract species (a = Neogoniolithon strictum, b = Jania adhaerens, c = Udotea luna, d = Halimeda scabra, e = CCA)
with the photosynthesis inhibitor DCMU in the light (500 μmol photons m− 2 s− 1) and dark at ambient and low pH. Means ± standard errors are shown. Different
lowercase letters represent significant differences from a two-way ANOVA with Light x pCO2 treatments (P < 0.05). *CCA calcification rates are normalized to g of flexible 2-D surface due to high 3-D complexity.

C. McNicholl and M.S. Koch

Journal of Experimental Marine Biology and Ecology 535 (2021) 151489

10

effect on FKRT species net calcification rates under high irradiance
(500 μmol photons m− 2 s− 1) except for U. luna. Neogoniolithon strictum
and J. adhaerens maintained 87% of their calcification rates and
H. scabra increased net calcification rates 16% at high light and low pH.
Two other Halimeda species (H. digitate and H. opuntia) from CO2 seep
sites in Indonesia at low pH (7.7) were also found to increase calcifi-
cation rates in the light (131% and 41%, respectively) relative to adja-
cent ambient pH control sites (Vogel et al. 2015a). The FKRT species that
maintained calcification rates under low pH and elevated pCO2 in the
light had a strong positive relationship between net photosynthesis and
calcification at ambient (R2 0.64 to 0.98) and low pH (R2 0.87 to 0.96).
While the importance of photosynthesis in macroalgal calcification has
long been appreciated (Borowitzka 1981; Koch et al. 2013; Pentecost
1978), how this relationship is sustained under low pH has not been
resolved. Further, in some species, as was shown for H. heteromorpha, net
photosynthetic rates do not always positively correspond to rates of
calcification at low pH (Brown et al. 2019). Photosynthesis has been
shown to elevate pH at the macroalgal thalli surface under low pH and
elevated pCO2 (Cornwall et al. 2015; Hofmann et al. 2016; McNicholl
et al. 2019). This photosynthetically-driven increase in surface pH can
mitigate the negative effects of bulk seawater acidification on net
calcification (Cornwall et al. 2014). Photosynthesis also elevates pH
within the calcifying space of corals and macroalgae, even under low
external bulk seawater pH (Comeau et al. 2018; Cornwall et al. 2017;
Venn et al. 2019).

In addition to photosynthesis, light-dependent proton pumps inde-
pendent of PSII may be important for macroalgal calcification. Three
species from the FKRT (H. scabra, N. strictum, and J. adhaerens) main-
tained 22% to 34% of their calcification rates in the light independent of
PSII. These same three species were observed to control thalli surface H+

light/dark dynamics independent of PSII (McNicholl et al. 2019). Thus,
active pH regulation independent of PSII may be linked to calcification

in macroalgae. Calcification rates in the light without PSII were not
sustained at low pH in the present study, even though H+ dynamics
seemingly continue at low pH (McNicholl et al. 2020). This was possibly
due to unsustainable H+ transport requirements, acid-base disfunction,
and/or changes in the electrochemical gradients of H+ across the plas-
malemma. Thus, we suggest photosynthesis and light-triggered proton
pumps may promote calcification in the light at ambient pH, but proton
pumps become overwhelmed under elevated [H+] at low pH. Under this
scenario, calcification would be more dependent on high rates of
photosynthesis as ocean pH declines.

Although the FKRT species were reliant on photosynthesis to facili-
tate high calcification rates, Halimeda species growing on reef walls
maintained similarly high calcification rates at 10-fold lower irradiance.
The distinct morphology and physiology of the three Halimeda species
examined in this study may explain the divergent responses to low pH.
H. scabra from the FKRT that sustained calcification in the light at low
pH has relatively high organic:inorganic carbon ratios (Peach et al.
2017b; Vroom et al. 2003) indicating a high photosynthetic capacity. In
contrast, species with lower organic:inorganic carbon ratios, H. goreauii
and H. copiosa, did not compensate for low pH via photosynthesis while
growing in low light (50 μmol photons m− 2 s− 1). Low-light adapted
Halimeda species likely have an alternative strategy to promote high
rates of calcification. One hypothesis is that short diffusive pathways
(<5 μm, Peach et al. 2017a) that connect their calcifying space to external bulk seawater allow for efficient export of H+, a byproduct of calcification that can lower internal pH and limit further calcification. The path-length may also facilitate diffusive uptake of Ca2+ and CO3

2−

into the calcifying space in support of calcification. It has been shown by
Peach et al. (2017a) that the shorter the diffusive path length the greater
the %CaCO3 in Halimeda species from LCI, including the species exam-
ined herein. A short path-length morphology in H. goreauii and
H. copiosa, combined with a low respiration rate due to a low organic:

Fig. 6. Net calcification rates (n = 4) of Little Cayman Island species (two chlorophyte Halimeda species: a = Halimeda goreauii, b = Halimeda copiosa, and two
rhodophyte species from the Peyssonneliaceae family: c = Peyr, d = Peyp) in low light (50 μmol photons m− 2 s− 1) and dark at ambient and low pH. Totals are daily net
calcification rates calculated by combining light and dark calcification rates at respective pH and normalized to 24 h. Means ± standard errors are shown. Different
lowercase letters represent significant differences from a two-way ANOVA with Light x pCO2 treatments and post-hoc analysis with Holm-Sidak (P < 0.05). Peys- sonneliaceae calcification rates were normalized to surface area of thalli lobes (cm2). Significant differences in daily calcification rates were determined by t-tests and shown with capital letters (P < 0.05, unless otherwise shown).

C. McNicholl and M.S. Koch

Journal of Experimental Marine Biology and Ecology 535 (2021) 151489

11

inorganic ratio, likely accounts for high calcification rates measured for
LCI Halimeda at ambient pH, regardless of low photosynthetic rates. Low
light adaptation, and potentially a greater dependence on proton pumps
associated with biotic control, was shown by H. goreauii’s and
H. copiosa’s ability to maintain 39% and 22% of their calcification rates
in the dark at ambient pH. However, low-light adapted morphology and
physiology that allow Halimeda species to be effective and dominant
calcifiers on deep reefs (Littler et al. 1985; Vroom et al. 2003), may also
enhance their vulnerability to low pH. For example, low-irradiance
Halimeda species are likely more dependent on proton pumps for calci-
fication. We observed a significant loss of non-PSII light-induced calci-
fication in FKRT species at low pH, potentially an indicator of a decline
in biotic control of calcification. While we (McNicholl et al. 2019) and
others (De Beer and Larkum 2001) observed Halimeda species to possess
proton pumps, the loss of proton pump function in regards to calcifi-
cation has not been examined in low-light species at low pH.

Two species from the high light FKRT site (U. luna and CCA) also had
significant declines in net calcification in the light at low pH. Net
dissolution occurred in these two species in the dark under both ambient
and low pH. The apparent lower resistance to dissolution and lower pH
in the light may correspond to the proximity of these species’ calcifi-
cation sites to bulk seawater. Udotea luna calcification occurs within
external sheaths along thalli filaments that are directly exposed to bulk
seawater (Bohm 1978). Udotea luna dark dissolution rates were also
found to be high in both ambient and low pH treatments in 45Ca ex-
periments (McNicholl et al. 2020). Further, microsensor experiments
(McNicholl et al. 2019) singled out U. luna amongst FKRT species as
having the least ability to raise pH at the thalli surface in response to
light at low pH, leading to the suggestion that it has weak biotic control.
Meyer et al. (2016) showed Udotea flabellum to also exhibit net disso-
lution in the dark under ambient pH. Further, U. flabellum had a 36%
lower net calcification rate under low pH in the light, relative to con-
trols. CCA calcification sites are also proximate to seawater because of
its prostrate form. This group is often recognized as being highly
vulnerable to declining pH and elevated pCO2 due to a high‑magnesium
CaCO3 structure (Ries 2011). Secondary calcification also occurs be-
tween filaments in coralline macroalgae which are more exposed to bulk
seawater than cell wall calcification, and thus less resistant to increased
[H+] (Cornwall et al. 2017; Hofmann et al. 2012). While we observed a
negative low pH effect on calcification rates of prostrate encrusting
rhodophytes in these short-term incubations, Peyssonneliaceae (Dutra
et al. 2015) and CCA (Kamenos et al. 2016; Martin et al. 2013b) have
exhibited more robust responses to elevated pCO2 in longer-term
studies. Kamenos et al. (2013) also detected molecular-level changes
in CCA carbonate minerals that were exposed to abrupt, but not slow,
treatments of low pH (7.77). These data correspond to recent results
demonstrating the ability of CCA to acclimate to lower pH over several
generations (Cornwall et al. 2020). Thus, mechanisms and species-
specific resistance/vulnerabilities to future changes in pH and carbon-
ate chemistry need further examination in both short and long-term
studies.

Based on our research, we propose that negative responses to
elevated pCO2 and low pH at the organismal level for calcifying tropical
macroalgae are primarily associated with effects on net calcification in
the dark for high-light species. This is likely attributable to greater
dissolution in the dark at low pH. Even with greater dissolution in the
dark at low pH, daily net calcification rates can be unaffected by low pH
because of high net calcification rates in the light and low overall
calcification rates in the dark. Low pH and elevated pCO2 effects on daily
calcification rates appear to be greatest in species that exhibit declines in
net calcification rates in the light. Further, our inhibition experiments
lead us to suggest that PSII-independent calcification mechanisms may
become overwhelmed at low pH with greater [H+] in the bulk seawater.
Thus, low-light species’ morphology and strategy that evolved to sustain
calcification without high rates of photosynthesis might make them
more vulnerable to greater [H+] in the bulk seawater. Light-driven

processes, including photosynthesis and/or H+ control, will be essen-
tial to sustain or enhance daytime calcification to offset nighttime
dissolution and maintain a net positive daily calcification rate. Thus,
macroalgae able to maintain high calcification rates in the light (high
and low irradiance) at low pH, and/or sustain strong biotic control with
high [H+] in the bulk seawater, are expected to dominate under global
change.

Supplementary data to this article can be found online at https://doi.
org/10.1016/j.jembe.2020.151489.

CRediT authorship contribution statement

C. McNicholl: Investigation, Formal analysis, Writing – original
draft, Visualization, Writing – review & editing, Validation, Software,
Methodology, Conceptualization. M.S. Koch: Writing – original draft,
Visualization, Writing – review & editing, Project administration, Su-
pervision, Resources, Validation, Methodology, Conceptualization.

Declaration of Competing Interest

The authors declare that they have no known competing financial
interests or personal relationships that could have appeared to influence
the work reported in this paper.

Acknowledgements

This research was funded by the National Science Foundation Ocean
Acidification Program-CRI-OA Grant #1416376. The authors would like
to thank Chris Johnson, Kimberly McFarlane, and the undergraduate
students that assisted in the field and lab. Dr. Carrie Manfrino is
recognized for her support in the field, and the Cayman Island Marine
Conservation Board and Department of the Environment for permitting
our LCI research. We also appreciate the anonymous reviewers and
editors that significantly improved the manuscript.

References

Adey, W.H., 1998. Coral reefs: algal structured and mediated ecosystems in shallow,
turbulent, alkaline waters. J. Phycol. 34, 393–406.

Adey, W.H., Halfar, J., Williams, B., 2013. The coralline genus Clathromorphum Foslie
emend. Adey. Smithson. Contrib. Mar. Sci. 40, 1–41.

Andersson, A.J., Kuffner, I.B., Mackenzie, F.T., Jokiel, P.L., Rodgers, K.S., Tan, A., 2009.
Net loss of CaCO3 from a subtropical calcifying community due to seawater
acidification: mesocosm-scale experimental evidence. Biogeosciences 6, 1811–1823.
https://doi.org/10.5194/bg-6-1811-2009.

Anthony, K.R.N., Kline, D.I., Diaz-Pulido, G., Dove, S., Hoegh-Guldberg, O., 2008. Ocean
acidification causes bleaching and productivity loss in coral reef builders. Proc. Natl.
Acad. Sci. 105, 17442–17446. https://doi.org/10.1073/pnas.0804478105.

Basso, D., 2012. Carbonate production by calcareous red algae and global change.
Geodiversitas 34, 13–33. https://doi.org/10.5252/g2012n1a2.

Bohm, L., 1978. Application of the 45Ca tracer method for determination of calcification
rates in calcareous algae: effect of calcium exchange and differential saturation of
algal calcium pools. Mar. Biol. 47, 9–14. https://doi.org/10.1007/bf00397013.

Borowitzka, M.A., 1981. Photosynthesis and calcification in the articulated coralline red
algae Amphiroa anceps and A. foliacea. Mar. Biol. 62, 17–23. https://doi.org/
10.1007/BF00396947.

Borowitzka, M.A., Larkum, A.W.D., 1976. Calcification in the green alga Halimeda: IV.
The action of metabolic inhibitors on photosynthesis and calcification. J. Exp. Bot.
27, 894–907. https://doi.org/10.1093/jxb/27.5.894.

Borowitzka, M.A., Larkum, A.W.D., 1987. Calcification in algae: mechanisms and the
role of metabolism. CRC. Crit. Rev. Plant Sci. 6, 1–45. https://doi.org/10.1080/
07352688709382246.

Brown, K.T., Bender-Champ, D., Kenyon, T.M., Rémond, C., Hoegh-Guldberg, O.,
Dove, S., 2019. Temporal effects of ocean warming and acidification on coral–algal
competition. Coral Reefs 38, 297–309. https://doi.org/10.1007/s00338-019-01775-
y.

Chisholm, J.R.M., 2000. Calcification by crustose coralline algae on the northern great
barrier reef. Australia. Limnol. Oceanogr. 45, 1476–1484. https://doi.org/10.2307/
2670432.

Chisholm, J.R.M., Gattuso, J., 1991. Validation of the alkalinity anomaly technique for
investigating calcification of photosynthesis in coral reef communities. Limnol.
Oceanogr. 36, 1232–1239. https://doi.org/10.4319/lo.1991.36.6.1232.

Chrachri, A., Hopkinson, B.M., Flynn, K., Brownlee, C., Wheeler, G.L., 2018. Dynamic
changes in carbonate chemistry in the microenvironment around single marine

C. McNicholl and M.S. Koch

https://doi.org/10.1016/j.jembe.2020.151489

https://doi.org/10.1016/j.jembe.2020.151489

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0005

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0005

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0010

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0010

https://doi.org/10.5194/bg-6-1811-2009

https://doi.org/10.1073/pnas.0804478105

https://doi.org/10.5252/g2012n1a2

https://doi.org/10.1007/bf003970

13

https://doi.org/10.1007/BF00396947

https://doi.org/10.1007/BF00396947

https://doi.org/10.1093/jxb/27.5.894

https://doi.org/10.1080/07352688709382246

https://doi.org/10.1080/07352688709382246

https://doi.org/10.1007/s00338-019-01775-y

https://doi.org/10.1007/s00338-019-01775-y

https://doi.org/10.2307/2670432

https://doi.org/10.2307/2670432

https://doi.org/10.4319/lo.1991.36.6.1232

Journal of Experimental Marine Biology and Ecology 535 (2021) 151489

12

phytoplankton cells. Nat. Commun. 1–12. https://doi.org/10.1038/s41467-017-
02426-y.

Comeau, S., Carpenter, R.C., Edmunds, P.J., 2012. Coral reef calcifiers buffer their
response to ocean acidification using both bicarbonate and carbonate. Proc. R. Soc. B
280, 1–8. https://doi.org/10.1098/rspb.2012.2374.

Comeau, S., Edmunds, P.J., Spindel, N.B., Carpenter, R.C., 2013. The responses of eight
coral reef calcifiers to increasing partial pressure of CO2 do not exhibit a tipping
point. Limnol. Oceanogr. 58, 388–398. https://doi.org/10.4319/lo.2013.58.1.0388.

Comeau, S., Carpenter, R., Edmunds, P., 2016. The effects of pCO2 on photosynthesis and
respiration of tropical scleractinian corals and calcified algae. ICES J. Mar. Sci.
https://doi.org/10.1093/icesjms/fsv267.

Comeau, S., Cornwall, C.E., McCulloch, M.T., 2017. Decoupling between the response of
coral calcifying fluid pH and calcification to ocean acidification. Sci. Rep. 7, 7573.
https://doi.org/10.1038/s41598-017-08003-z.

Comeau, S., Cornwall, C.E., DeCarlo, T.M., Krieger, E., McCulloch, M.T., 2018. Similar
controls on calcification under ocean acidification across unrelated coral reef taxa.
Glob. Chang. Biol. 24, 1–12. https://doi.org/10.1111/gcb.14379.

Comeau, S., Cornwall, C.E., DeCarlo, T.M., Doo, S.S., Carpenter, R.C., McCulloch, M.T.,
2019. Resistance to ocean acidification in coral reef taxa is not gained by
acclimatization. Nat. Clim. Chang. 9, 477–483. https://doi.org/10.1038/s41558-
019-0486-9.

Cornwall, C.E., Hepburn, C.D., Pilditch, C.A., Hurd, C.L., 2013. Concentration boundary
layers around complex assemblages of macroalgae: implications for the effects of
ocean acidification on understory coralline algae. Limnol. Oceanogr. 58, 121–130.
https://doi.org/10.4319/lo.2013.58.1.0121.

Cornwall, C.E., Boyd, P.W., McGraw, C.M., Hepburn, C.D., Pilditch, C.A., Morris, J.N.,
Smith, A.M., Hurd, C.L., 2014. Diffusion boundary layers ameliorate the negative
effects of ocean acidification on the temperate coralline macroalga Arthrocardia
corymbosa. PLoS One 9, 1–9. https://doi.org/10.1371/journal.pone.0097235.

Cornwall, C.E., Pilditch, C.A., Hepburn, C.D., Hurd, C.L., 2015. Canopy macroalgae
influence understorey corallines’ metabolic control of near-surface pH and oxygen
concentration. Mar. Ecol. Prog. Ser. 525, 81–95. https://doi.org/10.3354/
meps11190.

Cornwall, C.E., Comeau, S., McCulloch, M.T., 2017. Coralline algae elevate pH at the site
of calcification under ocean acidification. Glob. Chang. Biol. 23, 4245–4256. https://
doi.org/10.1111/gcb.13673.

Cornwall, C.E., Comeau, S., DeCarlo, T.M., Larcombe, E., Moore, B., Giltrow, K.,
Puerzer, F., D’Alexis, Q., McCulloch, M.T., 2020. A coralline alga gains tolerance to
ocean acidification over multiple generations of exposure. Nat. Clim. Chang. 10,
143–146. https://doi.org/10.1038/s41558-019-0681-8.

Cyronak, T., Schulz, K.G., Jokiel, P.L., 2015. The omega myth : what really drives lower
calcification rates in an acidifying ocean. ICES J. Mar. Sci. 1–5. https://doi.org/
10.1093/icesjms/fsv075.

De Beer, D., Larkum, A.W.D., 2001. Photosynthesis and calcification in the calcifying
algae Halimeda discoidea studied with microsensors. Plant Cell Environ. 24,
1209–1217. https://doi.org/10.1046/j.1365-3040.2001.00772.x.

Diaz-Pulido, G., McCook, L.J., Larkum, A.D., Lotze, H.K., Raven, J.A., Schaffelke, B.,
Smith, J.E., Steneck, R.S., 2007. Vulnerability of Macroalgae of the Great Barrier
Reef to Climate Change, in: Climate Change and the Great Barrier Reef: A
Vulnerability Assessment.

Diaz-Pulido, G., Nash, M.C., Anthony, K.R.N., Bender, D., Opdyke, B.N., Reyes-Nivia, C.,
Troitzsch, U., 2014. Greenhouse conditions induce mineralogical changes and
dolomite accumulation in coralline algae on tropical reefs. Nat. Commun. 5, 1–9.
https://doi.org/10.1038/ncomms4310.

Dutra, E., Koch, M., Peach, K.E., Manfrino, C., 2015. Tropical crustose coralline algal
individual and community responses to elevated pCO2 under high and low
irradiance. ICES J. Mar. Sci. 707–721.

El Haïkali, B., Bensoussan, N., Romano, J.C., Bousquet, V., 2004. Estimation of
photosynthesis and calcification rates of Corallina elongata Ellis and Solander, 1786,
by measurements of dissolved oxygen, pH and total alkalinity. Sci. Mar. 68, 45–56.
https://doi.org/10.3989/scimar.2004.68n145.

Fabry, V.J., Seibel, B., Feely, R., Orr, J., 2008. Impacts of ocean acidification on marine
fauna and ecosystem processes. Int. Counc. Explor. Sea 414–432. https://doi.org/
10.1093/icesjms/fsn048.

Gao, K., Aruga, Y., Asada, K., Ishihara, T., Akano, T., Kiyohara, M., 1993. Calcification in
the articulated coralline alga Corallina pilulifera, with special reference to the effect
of elevated CO2 concentration. Mar. Biol. 117, 129–132. https://doi.org/10.1007/
BF00346434.

Gehlen, M., Séférian, R., Jones, D.O.B., Roy, T., Roth, R., Barry, J., Bopp, L., Doney, S.C.,
Dunne, J.P., Heinze, C., Joos, F., Orr, J.C., Resplandy, L., Segschneider, J.,
Tjiputra, J., 2014. Projected pH reductions by 2100 might put deep North Atlantic
biodiversity at risk. Biogeosciences 11, 6955–6967. https://doi.org/10.5194/bg-11-
6955-2014.

Hartin, C.A., Bond-Lamberty, B., Patel, P., Mundra, A., 2016. Ocean acidification over
the next three centuries using a simple global climate carbon-cycle model:
projections and sensitivities. Biogeosciences 13, 4329–4342. https://doi.org/
10.5194/bg-13-4329-2016.

Hoegh-Guldberg, O., Mumby, P.J., Hooten, A.J., Steneck, R.S., Greenfield, P., Gomez, E.,
Harvell, C.D., Sale, P.F., Edwards, A.J., Caldeira, K., Knowlton, N., Eakin, C.M.,
Iglesiaia-Prieto, R., Muthiga, N., Bradbury, R.H., Dubi, A., Hatziolos, M.E., 2007.
Coral reefs under rapid climate change and ocean acidification. Science 318,
1737–1742. https://doi.org/10.1126/science.1152509.

Hofmann, L.C., Bischof, K., 2014. Ocean acidification effects on calcifying macroalgae.
Aquat. Biol. 22, 261–279. https://doi.org/10.3354/ab00581.

Hofmann, L.C., Yildiz, G., Hanelt, D., Bischof, K., 2012. Physiological responses of the
calcifying rhodophyte, Corallina officinalis (L.), to future CO2 levels. Mar. Biol. 159,
783–792. https://doi.org/10.1007/s00227-011-1854-9.

Hofmann, L.C., Heiden, J., Bischof, K., Teichberg, M., 2014. Nutrient availability affects
the response of the calcifying chlorophyte Halimeda opuntia (L.) J.V. Lamouroux to
low pH. Planta 239, 231–242. https://doi.org/10.1007/s00425-013-1982-1.

Hofmann, L.C., Koch, M., De Beer, D., 2016. Biotic control of surface pH and evidence of
light-induced H+ pumping and Ca2+ -H+ exchange in a tropical crustose coralline
alga. PLoS One 11, 1–24. https://doi.org/10.1371/journal.pone.0159057.

Jokiel, P.L., 2011. Ocean acidification and control of reef coral calcification by boundary
layer limitation of proton flux. Bull. Mar. Sci. 87, 639–657. https://doi.org/
10.5343/bms.2010.1107.

Kamenos, N.A., Law, A., 2010. Temperature controls on coralline algal skeletal growth.
J. Phycol. 46, 331–335. https://doi.org/10.1111/j.1529-8817.2009.00780.x.

Kamenos, N.A., Cusack, M., Huthwelker, T., Lagarde, P., Scheibling, R.E., 2009. Mg-
lattice associations in red coralline algae. Geochim. Cosmochim. Acta 73,
1901–1907. https://doi.org/10.1016/j.gca.2009.01.010.

Kamenos, N.A., Burdett, H.L., Aloisio, E., Findlay, H.S., Martin, S., Longbone, C.,
Dunn, J., Widdicombe, S., Calosi, P., 2013. Coralline algal structure is more sensitive
to rate, rather than the magnitude, of ocean acidification. Glob. Chang. Biol. 19,
3621–3628. https://doi.org/10.1111/gcb.12351.

Kamenos, N.A., Perna, G., Gambi, M.C., Micheli, F., Kroeker, K.J., 2016. Coralline algae
in a naturally acidified ecosystem persist by maintaining control of skeletal
mineralogy and size. Proc. R. Soc. B Biol. Sci. 283, 20161159 https://doi.org/
10.1098/rspb.2016.1159.

Kato, A., Hikami, M., Kumagai, N.H., Suzuki, A., Nojiri, Y., Sakai, K., 2014. Negative
effects of ocean acidification on two crustose coralline species using genetically
homogeneous samples. Mar. Environ. Res. 94, 1–6. https://doi.org/10.1016/j.
marenvres.2013.10.010.

Koch, M., Bowes, G., Ross, C., Zhang, X.-H., 2013. Climate change and ocean
acidification effects on seagrasses and marine macroalgae. Glob. Chang. Biol. 19
https://doi.org/10.1111/j.1365-2486.2012.02791.x.

Kuffner, I.B., Lidz, B.H., Hudson, J.H., Anderson, J.S., 2015. A century of ocean warming
on Florida keys coral reefs: historic in situ observations. Estuar. Coasts 38,
1085–1096. https://doi.org/10.1007/s12237-014-9875-5.

Langdon, C., Gattuso, J., Andersson, A., 2010. Part 3: measurements of calcification and
dissolution of benthic organisms and communities. In: Riebesell, U., Fabry, V.J.,
Hansson, L. (Eds.), Guide to Best Practices for Ocean Acidification Research and Data
Reporting. European Commission, Office CDMA 03/115, B-1049, Brussels,
pp. 213–232.

Littler, M.M., Littler, D.S., Blair, S.M., Norris, J.N., 1985. Deepest known plant life
discovered on an uncharted seamount. Science 227, 57–59. https://doi.org/
10.1126/science.227.4682.57.

Marsh, J.A., 1970. Primary productivity of reef-building calcareous red algae. Ecology
51, 255–263.

Martin, S., Charnoz, A., Gattuso, J.P., 2013a. Photosynthesis, respiration and
calcification in the Mediterranean crustose coralline alga Lithophyllum cabiochae
(Corallinales, Rhodophyta). Eur. J. Phycol. 48, 163–172. https://doi.org/10.1080/
09670262.2013.786790.

Martin, S., Cohu, S., Vignot, C., Zimmerman, G., Gattuso, J.P., 2013b. One-year
experiment on the physiological response of the Mediterranean crustose coralline
alga, Lithophyllum cabiochae, to elevated pCO2 and temperature. Ecol. Evol. 3,
676–693. https://doi.org/10.1002/ece3.475.

Mccoy, S.J., Kamenos, N.A., 2015. Coralline algae (Rhodophyta) in a changing world:
integrating ecological, physiological, and geochemical responses to global change.
J. Phycol. 51, 6–24. https://doi.org/10.1111/jpy.12262.

McNicholl, C., Koch, M.S., Hofmann, L.C., 2019. Photosynthesis and light-dependent
proton pumps increase boundary layer pH in tropical macroalgae: a proposed
mechanism to sustain calcification under ocean acidification. J. Exp. Mar. Bio. Ecol.
521 (1–12).

McNicholl, C., Koch, M.S., Swarzenski, P.W., Oberhaensli, F.R., Taylor, A., Batista, M.G.,
Metian, M., 2020. Ocean acidification effects on calcification and dissolution in
tropical reef macroalgae. Coral Reefs. https://doi.org/10.1007/s00338-020-01991-
x.

Meyer, F.W., Schubert, N., Diele, K., Teichberg, M., Wild, C., Enriquez, S., 2016. Effect of
inorganic and organic carbon enrichments (DIC and DOC) on the photosynthesis and
calcification rates of two calcifying green algae from a Caribbean reef lagoon. PLoS
One 11. https://doi.org/10.1371/journal.pone.0160268.

Nelson, W.A., 2009. Calcified macroalgae–critical to coastal ecosystems and vulnerable
to change: a review. Mar. Freshw. Res. 60, 787–801. https://doi.org/10.1071/
MF08335.

Noisette, F., Duong, G., Six, C., Davoult, D., Martin, S., 2013. Effects of elevated pCO2 on
the metabolism of a temperate rhodolith Lithothamnion corallioides grown under
different temperatures. J. Phycol. 49, 746–757. https://doi.org/10.1111/jpy.12085.

Orr, J.C., Fabry, V.J., Aumont, O., Bopp, L., Doney, S.C., Feely, R.A., Gnanadesikan, A.,
Gruber, N., Ishida, A., Joos, F., Key, R.M., Lindsay, K., Maier-Reimer, E., Matear, R.,
Monfray, P., Mouchet, A., Najjar, R.G., Plattner, G.-K., Rodgers, K.B., Sabine, C.L.,
Sarmiento, J.L., Schlitzer, R., Slater, R.D., Totterdell, I.J., Weirig, M.-F.,
Yamanaka, Y., Yool, A., 2005. Anthropogenic Ocean acidification over the twenty-
first century and its impact on calcifying organisms. Nature 437, 681–686. https://
doi.org/10.1038/nature04095.

Peach, K.E., Koch, M.S., Blackwelder, P.L., 2016. Effects of elevated pCO2 and irradiance
on growth, photosynthesis and calcification in Halimeda discoidea. Mar. Ecol. Prog.
Ser. 544, 143–158. https://doi.org/10.3354/meps11591.

C. McNicholl and M.S. Koch

https://doi.org/10.1038/s41467-017-02426-y

https://doi.org/10.1038/s41467-017-02426-y

https://doi.org/10.1098/rspb.2012.2374

https://doi.org/10.4319/lo.2013.58.1.0388

https://doi.org/10.1093/icesjms/fsv267

https://doi.org/10.1038/s41598-017-08003-z

https://doi.org/10.1111/gcb.14379

https://doi.org/10.1038/s41558-019-0486-9

https://doi.org/10.1038/s41558-019-0486-9

https://doi.org/10.4319/lo.2013.58.1.0121

https://doi.org/10.1371/journal.pone.0097235

https://doi.org/10.3354/meps11190

https://doi.org/10.3354/meps11190

https://doi.org/10.1111/gcb.13673

https://doi.org/10.1111/gcb.13673

https://doi.org/10.1038/s41558-019-0681-8

https://doi.org/10.1093/icesjms/fsv075

https://doi.org/10.1093/icesjms/fsv075

https://doi.org/10.1046/j.1365-3040.2001.00772.x

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0135

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0135

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0135

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0135

https://doi.org/10.1038/ncomms4310

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0145

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0145

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0145

https://doi.org/10.3989/scimar.2004.68n145

https://doi.org/10.1093/icesjms/fsn048

https://doi.org/10.1093/icesjms/fsn048

https://doi.org/10.1007/BF00346434

https://doi.org/10.1007/BF00346434

https://doi.org/10.5194/bg-11-6955-2014

https://doi.org/10.5194/bg-11-6955-2014

https://doi.org/10.5194/bg-13-4329-2016

https://doi.org/10.5194/bg-13-4329-2016

https://doi.org/10.1126/science.1152509

https://doi.org/10.3354/ab00581

https://doi.org/10.1007/s00227-011-1854-9

https://doi.org/10.1007/s00425-013-1982-1

https://doi.org/10.1371/journal.pone.0159057

https://doi.org/10.5343/bms.2010.1107

https://doi.org/10.5343/bms.2010.1107

https://doi.org/10.1111/j.1529-8817.2009.00780.x

https://doi.org/10.1016/j.gca.2009.01.010

https://doi.org/10.1111/gcb.12351

https://doi.org/10.1098/rspb.2016.1159

https://doi.org/10.1098/rspb.2016.1159

https://doi.org/10.1016/j.marenvres.2013.10.010

https://doi.org/10.1016/j.marenvres.2013.10.010

https://doi.org/10.1111/j.1365-2486.2012.02791.x

https://doi.org/10.1007/s12237-014-9875-5

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0240

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0240

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0240

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0240

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0240

https://doi.org/10.1126/science.227.4682.57

https://doi.org/10.1126/science.227.4682.57

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0250

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0250

https://doi.org/10.1080/09670262.2013.786790

https://doi.org/10.1080/09670262.2013.786790

https://doi.org/10.1002/ece3.475

https://doi.org/10.1111/jpy.12262

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0270

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0270

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0270

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0270

https://doi.org/10.1007/s00338-020-01991-x

https://doi.org/10.1007/s00338-020-01991-x

https://doi.org/10.1371/journal.pone.0160268

https://doi.org/10.1071/MF08335

https://doi.org/10.1071/MF08335

https://doi.org/10.1111/jpy.12085

https://doi.org/10.1038/nature04095

https://doi.org/10.1038/nature04095

https://doi.org/10.3354/meps11591

Journal of Experimental Marine Biology and Ecology 535 (2021) 151489
13

Peach, K.E., Koch, M.S., Blackwelder, P.L., Guerrero-Given, D., Kamasawa, N., 2017a.
Primary utricle structure of six Halimeda species and potential relevance for ocean
acidification tolerance. Bot. Mar. 60, 1–11. https://doi.org/10.1515/bot-2016-0055.

Peach, K.E., Koch, M.S., Blackwelder, P.L., Manfrino, C., 2017b. Calcification and
photophysiology responses to elevated pCO2 in six Halimeda species from contrasting
irradiance environments on little Cayman island reefs. J. Exp. Mar. Bio. Ecol. 486,
114–126. https://doi.org/10.1016/j.jembe.2016.09.008.

Pentecost, A., 1978. Calcification and photosynthesis in Corallina officinalis l. using the
14CO2 method. Br. Phycol. J. 13, 383–390. https://doi.org/10.1080/
00071617800650431.

Pierrot, D., Lewis, E., Wallace, D., 2006. MS Excel Program Developed for CO2 Systems
Calculations: ORNL/CDIAC 105a.

Porzio, L., Buia, M.C., Hall-Spencer, J.M., 2011. Effects of ocean acidification on
macroalgal communities. J. Exp. Mar. Bio. Ecol. 400, 278–287. https://doi.org/
10.1016/j.jembe.2011.02.011.

Price, N.N., Hamilton, S.L., Tootell, J.S., Smith, J.E., 2011. Species-specific consequences
of ocean acidification for the calcareous tropical green algae Halimeda. Mar. Ecol.
Prog. Ser. 440, 67–78. https://doi.org/10.3354/meps09309.

Pueschel, C.M., Judson, B.L., Wegeberg, S., 2005. Decalcification during epithallial cell
turnover in Jania adhaerens (Corallinales, Rhodophyta). Phycologia 44, 156–162.
https://doi.org/10.2216/0031-8884(2005)44[156:DDECTI]2.0.CO;2.

Raven, J.A., Hurd, C.L., 2012. Ecophysiology of photosynthesis in macroalgae.
Photosynth. Res. 113, 105–125. https://doi.org/10.1007/s11120-012-9768-z.

Ries, J.B., 2011. Skeletal mineralogy in a high-CO2 world. J. Exp. Mar. Bio. Ecol. 403,
54–64. https://doi.org/10.1016/j.jembe.2011.04.006.

Ries, J.B., Cohen, A.L., McCorkle, D.C., 2009. Marine calcifiers exhibit mixed responses
to CO2-induced ocean acidification. Geology 37, 1131–1134. https://doi.org/
10.1130/G30210A.1.

Semesi, I.S., Kangwe, J., Björk, M., 2009. Alterations in seawater pH and CO2 affect
calcification and photosynthesis in the tropical coralline alga, Hydrolithon sp.

(Rhodophyta). Estuar. Coast. Shelf Sci. 84, 337–341. https://doi.org/10.1016/j.
ecss.2009.03.038.

Taylor, A.R., Brownlee, C., Wheeler, G.L., 2012. Proton channels in algae: reasons to be
excited. Trends Plant Sci. 17, 675–684. https://doi.org/10.1016/j.
tplants.2012.06.009.

Venn, A.A., Tambutté, E., Caminiti-Segonds, N., Techer, N., Allemand, D., Tambutté, S.,
2019. Effects of light and darkness on pH regulation in three coral species exposed to
seawater acidification. Sci. Rep. 9, 1–12. https://doi.org/10.1038/s41598-018-
38168-0.

Vogel, N., Fabricius, K.S., Strahl, J., Noonan, S., Wild, C., Uthicke, S., 2015a. Calcareous
green alga Halimeda tolerates ocean acidification conditions at tropical carbon
dioxide seeps. Limnol. Oceanogr. 60, 263–275. https://doi.org/10.1002/lno.10021.

Vogel, N., Meyer, F.W., Wild, C., Uthicke, S., 2015b. Decreased light availability can
amplify negative impacts of ocean acidification on calcifying coral reef organisms.
Mar. Ecol. Prog. Ser. 521, 49–61. https://doi.org/10.3354/meps11088.

Vroom, P.S., Smith, C.M., Coyer, J.A., Walters, L.J., Hunter, C.L., Beach, K.S., Smith, J.E.,
2003. Field biology of Halimeda tuna (Bryopsidales, Chlorophyta) across a depth
gradient: comparative growth, survivorship, recruitment, and reproduction.
Hydrobiologia 501, 149–166. https://doi.org/10.1023/A:1026287816324.

Wizemann, A., Meyer, F.W., Westphal, H., 2014. A new model for the calcification of the
green macro-alga Halimeda opuntia (Lamouroux). Coral Reefs 33, 951–964. https://
doi.org/10.1007/s00338-014-1183-9.

Zou, D., Gao, K., Luo, H., 2011. Short- and long-term effects of elevated CO2 on
photosynthesis and respiration in the marine macroalga Hizikia Fusiformis
(Sargassaceae, Phaeophyta) grown at low and high N supplies. J. Phycol. 47, 87–97.
https://doi.org/10.1111/j.1529-8817.2010.00929.x.

Zweng, R.C., Koch, M.S., Bowes, G., 2018. The role of irradiance and C-use strategies in
tropical macroalgae photosynthetic response to ocean acidification. Sci. Rep. 8,
1–11. https://doi.org/10.1038/s41598-018-27333-0.

C. McNicholl and M.S. Koch

https://doi.org/10.1515/bot-2016-0055

https://doi.org/10.1016/j.jembe.2016.09.008

https://doi.org/10.1080/00071617800650431

https://doi.org/10.1080/00071617800650431

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0320

http://refhub.elsevier.com/S0022-0981(19)30489-7/rf0320

https://doi.org/10.1016/j.jembe.2011.02.011

https://doi.org/10.1016/j.jembe.2011.02.011

https://doi.org/10.3354/meps09309

https://doi.org/10.2216/0031-8884(2005)44[156:DDECTI]2.0.CO;2

https://doi.org/10.1007/s11120-012-9768-z

https://doi.org/10.1016/j.jembe.2011.04.006

https://doi.org/10.1130/G30210A.1

https://doi.org/10.1130/G30210A.1

https://doi.org/10.1016/j.ecss.2009.03.038

https://doi.org/10.1016/j.ecss.2009.03.038

https://doi.org/10.1016/j.tplants.2012.06.009

https://doi.org/10.1016/j.tplants.2012.06.009

https://doi.org/10.1038/s41598-018-38168-0

https://doi.org/10.1038/s41598-018-38168-0

https://doi.org/10.1002/lno.10021

https://doi.org/10.3354/meps11088

https://doi.org/10.1023/A:1026287816324

https://doi.org/10.1007/s00338-014-1183-9

https://doi.org/10.1007/s00338-014-1183-9

https://doi.org/10.1111/j.1529-8817.2010.00929.x

https://doi.org/10.1038/s41598-018-27333-0

• Irradiance, photosynthesis and elevated pCO2 effects on net calcification in tropical reef macroalgae

1 Introduction
2 Materials & methods
2.1 Calcified macroalgae collection and site parameters
2.1.1 Florida keys reef tract (high light environment)
2.1.2 Little Cayman Island reef (low light environment)
2.2 Elevated pCO2 experiments
2.3 PSII inhibitor experiments
2.4 Calculation of net calcification and chemistry analysis
2.5 Photosynthesis and respiration
2.6 Statistical analysis
3 Results
3.1 Carbonate chemistry and △pH, △TA and △O2
3.2 High light species (FKRT) calcification
3.2.1 Photosynthesis and calcification
3.2.2 Inhibitor experiments (FKRT)
3.3 Low light species (LCI) calcification
4 Discussion
CRediT authorship contribution statement
Declaration of Competing Interest
Acknowledgements
References

Glob Change Biol. 2020;00:1–14. wileyonlinelibrary.com/journal/gcb  |  1© 2020 John Wiley & Sons Ltd

Received: 28 July 2020  |  Accepted: 13 November 2020
DOI: 10.1111/gcb.15455

P R I M A R Y R E S E A R C H A R T I C L E

Ocean acidification locks algal communities in a species-poor
early successional stage

Ben P. Harvey1  | Koetsu Kon1  | Sylvain Agostini1  | Shigeki Wada1  |
Jason M. Hall-Spencer1,2

1Shimoda Marine Research Center,
University of Tsukuba, Shizuoka, Japan
2Marine Biology and Ecology Research
Centre, University of Plymouth, Plymouth,
UK

Correspondence
Ben P. Harvey, Shimoda Marine Research
Center, University of Tsukuba, 5-10-1
Shimoda, Shizuoka 415-0025, Japan.
Email: ben.harvey@shimoda.tsukuba.ac.jp

Funding information
Japan Society for the Promotion
of Science, Grant/Award Number:
17K17622; Ministry of Environment,
Government of Japan, Grant/Award
Number: 4RF-1701; University of Tsukuba

Abstract
Long-term exposure to CO2-enriched waters can considerably alter marine biological
community development, often resulting in simplified systems dominated by turf
algae that possess reduced biodiversity and low ecological complexity. Current un-
derstanding of the underlying processes by which ocean acidification alters biologi-
cal community development and stability remains limited, making the management
of such shifts problematic. Here, we deployed recruitment tiles in reference (pHT
8.137 ± 0.056 SD) and CO2-enriched conditions (pHT 7.788 ± 0.105 SD) at a volcanic
CO2 seep in Japan to assess the underlying processes and patterns of algal commu-
nity development. We assessed (i) algal community succession in two different sea-
sons (Cooler months: January–July, and warmer months: July–January), (ii) the effects
of initial community composition on subsequent community succession (by recipro-
cally transplanting preestablished communities for a further 6 months), and (iii) the
community production of resulting communities, to assess how their functioning was
altered (following 12 months recruitment). Settlement tiles became dominated by
turf algae under CO2-enrichment and had lower biomass, diversity and complexity, a
pattern consistent across seasons. This locked the community in a species-poor early
successional stage. In terms of community functioning, the elevated pCO2 commu-
nity had greater net community production, but this did not result in increased algal
community cover, biomass, biodiversity or structural complexity. Taken together, this
shows that both new and established communities become simplified by rising CO2
levels. Our transplant of preestablished communities from enriched CO2 to refer-
ence conditions demonstrated their high resilience, since they became indistinguish-
able from communities maintained entirely in reference conditions. This shows that
meaningful reductions in pCO2 can enable the recovery of algal communities. By
understanding the ecological processes responsible for driving shifts in community
composition, we can better assess how communities are likely to be altered by ocean
acidification.

K E Y W O R D S
CO2 seeps, community dynamics, competition, ecosystem function, global change ecology,
inhibition, turf algae

www.wileyonlinelibrary.com/journal/gcb

mailto:

https://orcid.org/0000-0002-4971-1634

https://orcid.org/0000-0003-1379-0702

https://orcid.org/0000-0001-9040-9296

https://orcid.org/0000-0001-6893-7498

https://orcid.org/0000-0002-6915-2518

mailto:ben.harvey@shimoda.tsukuba.ac.jp

http://crossmark.crossref.org/dialog/?doi=10.1111%2Fgcb.15455&domain=pdf&date_stamp=2021-01-10

2  |    HARVEY Et Al.

1   |   I N T R O D U C T I O N

The oceanic uptake of anthropogenic carbon dioxide emissions is
a global environmental issue termed ocean acidification. The ef-
fects of ocean acidification are detrimental to a wide range of ma-
rine organisms (Harvey et al., 2013; Kroeker, Kordas, et al., 2013),
and this affects ecosystem functioning and the goods and services
that people derive from marine resources (Gattuso et al., 2015;
Hall-Spencer & Harvey, 2019). To better understand the effects
of ocean acidification, there has been an effort in recent years to
move beyond aquarium-based experiments on single species to-
wards in-situ experiments (e.g. Albright et al., 2016, 2018; Brown
et al., 2016), long-term mesocosm observations (e.g. Algueró-Muñiz
et al., 2017; Moulin et al., 2015), and studies using natural CO2 seeps
(e.g. Agostini et al., 2018; Fabricius et al., 2011; Hall-Spencer et al.,
2008; Milazzo et al., 2014). These approaches have shown that long-
term exposure to ocean acidification conditions projected for the
end of the century fundamentally alters the composition of marine
biological communities, usually resulting in simplified systems with
reduced biodiversity and less ecological complexity (Agostini et al.,
2018; Sunday et al., 2017; Vizzini et al., 2017). Many of these studies
have been observation-based, and so an understanding of the un-
derlying processes responsible for driving these patterns in commu-
nity development remains limited. To help assess the future effects
of ocean acidification, it would be useful to better understand how
community development processes are affected by rising levels of
seawater CO2 (Gaylord et al., 2015), and how such changes will influ-
ence their associated ecosystem functioning.

Ecological theory suggests that the successional trajectories of
‘disturbed’ marine subtidal communities will be primarily driven by
physical stresses, competition for resources through the mecha-
nisms of ‘facilitation’ and ‘inhibition’ (Connell & Slatyer, 1977) and
the strength of associated bottom-up and top-down interactions
(Gruner et al., 2008; Jenkins et al., 1999). One of the difficulties in
predicting how community development will be affected by ocean
acidification is that the changes in carbonate chemistry can simul-
taneously act as both resource and stressor (Connell et al., 2013,
2018; Milazzo et al., 2019). It provides a bottom-up resource to
primary producers by enhancing the availability of bicarbonate and
CO2 (Connell et al., 2013; Koch et al., 2013), but also acts as a phys-
ical stressor to many organisms (including calcified primary produc-
ers) via negative effects on their physiology (Harvey et al., 2013;
Kroeker, Kordas, et al., 2013). Subsequently, marine communities are
expected to be re-organized by the effects of ocean acidification.
Ocean acidification alters the initial successional trajectories of algal
communities, which lead to dominance by fleshy algae over calcified
algae in acidified conditions projected for the end of the century in
both temperate and tropical settings (Crook et al., 2016; Kroeker
et al., 2012). Enriched CO2 alters competitive interactions, acting as
a physical stressor to calcified macroalgae whereas turf algae can
use the additional carbon to boost growth, which allows turf algae to
attain dominance (also see Connell et al., 2018). Fast-growing oppor-
tunistic (r-selected) turf algal species are usually suppressed beneath

macroalgal canopies on temperate reefs (Johnson & Mann, 1988)
and by top-down control of grazers in coral reefs (Hughes et al.,
2007). In the absence of strong competition or compensatory pro-
cesses (e.g. Connell et al., 2018; Ghedini & Connell, 2016; Ghedini
et al., 2015), turf species can become dominant thereby changing
the ecosystem state.

Under present-day conditions, it has been suggested that despite
bottom-up control of primary production being pervasive, top-down
control by consumers has a stronger influence on the trajectories of
algal community succession (Gruner et al., 2008; Hillebrand et al.,
2007). For example, intense grazing by sea urchins and herbivo-
rous fish can prevent kelp forest growth resulting in ‘urchin barrens’
dominated by crustose coralline algae (Kelly et al., 2016; Ling et al.,
2015). Ocean acidification is expected to reduce bottom-up control
on those species which are carbon-limited, as long as sufficient nu-
trients are available (Celis-Plá et al., 2015; Gordillo et al., 2003; Li
et al., 2012). Top-down control by benthic invertebrates in acidified
conditions may also diminish, given that at CO2 seeps the abun-
dance and size of many marine fauna are reduced (Garilli et al., 2015;
Harvey et al., 2016, 2018), with such examples as the observed num-
ber of sea urchin feeding halos being reduced in a CO2 seep (Kroeker
et al., 2013). Fish communities include a greater proportion of her-
bivorous fish within acidified conditions (during the period of peak
macroalgae biomass; Cattano et al., 2020), and so it may be possible
in some systems for fish to maintain top-down control (Baggini et al.,
2015). Taken together, any strong reductions in bottom-up and/or
top-down control are likely to alter community successional tra-
jectories and allow r-selected opportunistic species to outcompete
other species and dominate under ocean acidification.

Seasonality is an important aspect of shallow-water ecosystems,
and yet the consequences of seasonally induced environmental fluc-
tuations have rarely been considered in ocean acidification studies
(Baggini et al., 2014; Godbold & Solan, 2013). Algal communities in
temperate and warm temperate ecosystems experience large seasonal
changes in environmental conditions (Figure 1), which result in con-
siderable temporal shifts with a period of high recruitment and peak
biomass typically occurring in late spring. Thus the responses of algal
communities to ocean acidification will likely be strongly influenced by
seasonality (Baggini et al., 2014). In the Northern Pacific Ocean, this
is further complicated by the occurrence of typhoons, which typically
occur between July and October in Japan. Typhoons act as a substan-
tial physical disturbance that affects benthic community structure and
habitat complexity, such as through the removal of corals (Done, 1992),
macroalgae (Cattano et al., 2020) and seagrass cover (Wilson et al.,
2020), and can indirectly change the community function of associated
species (e.g. fish; Cattano et al., 2020).

Observations at natural CO2 seeps worldwide provide a good
understanding of how long-term ocean acidification simplifies the
composition of climax communities (Foo et al., 2018; González-
Delgado & Hernández, 2018; Hall-Spencer & Harvey, 2019), yet it
remains unclear whether these simplified communities develop due
to altered successional trajectory, stunted community development
(via successional inhibition) or are driven by reduced bottom-up and/

    |  3HARVEY Et Al.

or top-down control. To address these gaps, we deployed recruit-
ment tiles in reference (~350 μatm pCO2) and acidified (~900 μatm
pCO2) conditions using a natural CO2 seep area as an analogue for
end of the century pCO2 conditions (the representative concentra-
tion pathway (RCP) 8.5 scenario, 851 to 1370 μatm; IPCC, 2013) to
assess the early to mid-successional trajectories of algal communities
in two different seasons (cooler months January to July, and warmer
months July to January). The study was carried out over these two
time periods to investigate whether the effects of ocean acidification
on community development are temporally consistent. Following this
we carried out a reciprocal transplant of some of those established
communities, in order to assess the effects of initial community com-
position on subsequent community succession in the reference and
acidified conditions. Finally, we assessed the community production
of these reciprocally transplanted communities (including the asso-
ciated sessile invertebrate communities which contribute in terms of
respiration), in order to determine how any changes in community
composition will alter their ecosystem functioning.

2   |   M AT E R I A L S A N D M E T H O D S

2.1  |  Experimental design

To investigate our core question of how ocean acidification in-
fluences early community succession of algal communities, ex-
periments using recruitment tiles were carried out using an
acidified area of the Shikine Island CO2 seep, Japan (34°19′9ʺN,
139°12′18ʺE), and a nearby reference pCO2 area in an adjacent bay

(~600 m away by the shortest route). Both the reference and acidi-
fied locations (hereafter ‘350 μatm’ and ‘900 μatm’, respectively)
have had their carbonate chemistry and biology well character-
ized previously (Agostini et al., 2015, 2018; Cattano et al., 2020;
Harvey et al., 2018, 2019; Kerfahi et al., 2020; Witkowski et al.,
2019), and we present 2 months of additional original pHT (Figure
S1) and temperature data collected at the ‘900 μatm’ location with
a Durafet sensor (SeaFET, Sea-Bird Scientific) using the same ap-
proach as Agostini et al. (2018). Salinity was measured concurrently
using Hobo conductivity loggers (U24-002-C), and discrete sam-
ples for total alkalinity were collected throughout the study period,
with total alkalinity measured using an auto-titrator (916 Ti-Touch,
Metrohm). In summary, the ‘350 μatm’ location had a mean pHT of
8.137 ± 0.056 (SD) and the ‘900 μatm’ location had a mean pHT of
7.781 ± 0.105 (SD), and the mean carbonate chemistry of the two
locations is presented in Table 1. Long-term temperature data were
recorded over a 1-year period by deploying a temperature logger
(HOBO Pendant Temperature/Light 64K Data Logger) at ~6 m
depth in each site. The ‘900 μatm’ elevated pCO2 location repre-
sents an end of the century projection for reductions in pH (the RCP
8.5 scenario; IPCC, 2013), and was not confounded by differences in
temperature, salinity, dissolved oxygen, total alkalinity, nutrients or
depth relative to reference sites used for comparison (Agostini et al.,
2015, 2018; Harvey et al., 2019). Our basalt recruitment tiles were
130 × 130 × 15 mm and were secured using individual anchor bolts
(8.5 mm width, 70 mm length) drilled into rock by SCUBA divers at
~6 m depth (Nemo Underwater Drill). The tiles at each location were
deployed haphazardly across a c. 400 m2 area (with at least 5 m be-
tween individual tiles), fixed to upward-facing substrata.

F I G U R E 1  Conceptual representation
of the recruitment tile treatments. (a)
Tiles were deployed for 6 months during
the ‘Cold Period’ or ‘Warm Period’ in
either reference pCO2 (350 μatm; blue
line) or acidified conditions (900 μatm;
red line). (b) Tiles from the ‘Warm Period’
were then used as part of a reciprocal
transplant either being transplanted into
350 or 900 μatm conditions for a further
6 months

4  |    HARVEY Et Al.

2.1.1  |  Seasonal experiment

For the first experiment, five recruitment tiles were individually de-
ployed in each location (350 and 900 μatm) during the cooler months
of January 2017–July 2017 (hereafter termed ‘Cold Period’), with
eight recruitment tiles deployed in each location (350 and 900 μatm)
during the warmer months of July 2017–January 2018 (hereafter
termed ‘Warm Period’). Mean seawater temperature (±SD) during
the ‘Cold Period’ was 18.14 ± 1.81°C at 350 μatm and 18.07 ± 1.63°C
at 900 μatm, and during the ‘Warm Period’ was 22.86 ± 2.97°C at
350 μatm and 22.67 ± 2.83°C at 900 μatm. See Figure 1a for a con-
ceptual overview of the experimental design.

2.1.2  |  Reciprocal experiment

For the second experiment, tiles from the ‘Warm Period’ of the
seasonal experiment were used (each tile had an algal community
following 6 months recruitment). Sixteen tiles from the seasonal ex-
periment were reciprocally transplanted into the 350 and 900 μatm
locations for a further 6 months to assess the effects of initial com-
munity composition on subsequent community succession in refer-
ence and acidified conditions (four tiles in each combination, see
Figure 1b).

2.2  |  Community analysis

For both the seasonal experiment and the reciprocal transplanta-
tion experiment, following the 6 month experimental period each
tile was brought into the laboratory and photographed (Nikon
D7200, Nikon). For each 12 month tile, two photos were taken to
image the upperstorey community, and (after removal of the up-
perstorey community by hand) the understorey community. These
data were then combined for analysis. Community composition
was assessed using ImageJ (Abràmoff et al., 2004) by overlaying
64 points on a grid, and recording the abundance of each algal
functional group underlying that point. Functional groups used
followed Steneck and Dethier (1994), with algal grouping being

based on their morphology, thallus size and complexity (micro-
algae, filamentous algae, foliose algae, corticated foliose algae,
corticated macrophytes, leathery macrophytes, articulated cal-
careous algae and crustose algae). The habitat complexity of each
tile was determined by combining the abundance of each algal
functional group with a rank (0–5) based on the biogenic habitat
complexity provided by that functional group. The ranking was
based on categories assigned by Steneck and Dethier (1994): mi-
croalgae = 1, filamentous algae = 2, foliose algae = 3, corticated
foliose algae = 3.5, corticated macrophytes = 4, leathery macro-
phytes = 5, articulated calcareous algae = 4 and crustose algae = 3.
The habitat complexity score was then normalized to between 0
and 1.

2.3  |  Community production

Community production and respiration of individual tiles were as-
sessed by measuring, with an Orion 4-Star pH and dissolved oxy-
gen meter (Thermo Scientific), the changes in dissolved oxygen
concentrations during an incubation within a 2.5 L seawater con-
tainer (15 cm wide × 20 cm length × 10 cm height) in a tempera-
ture-controlled water bath. Magnetic stirrers (M-1 Controller and
MS101A Stirrer, AS-One) were used to continuously mix the sea-
water within each container throughout measurements. Seawater
for each treatment (pHNBS 8.05 ± 0.01 SD vs. pHNBS 7.83 ± 0.01
SD) was collected from the same location off-shore (oceanic pHNBS
8.05) and treatments of pHNBS 7.83 ± 0.01 SD were acquired by
bubbling with pure CO2 (Fukurow pH Controller; Aqua Geek).
Community production and respiration were measured over a
150 min period; first determining oxygen production (60 min light
period, c. 200 μmol m−2 s−1), and after a 30 min dark period, oxygen
consumption (60 min dark period). Net community production and
community respiration were measured during the light and dark
periods, respectively, with gross primary production calculated as
the net community production minus community respiration. All
three measurements are presented as mmol O2 h

−1 m−2. To assess
whether any changes in respiration rates were driven by an altered
sessile invertebrate community for the 12 month communities, the

T A B L E 1  Summary of the carbonate chemistry for the 350 and 900 μatm locations. The pHT (350 μatm, n = 1964; 900 μatm, n = 10,818),
salinity (350 μatm, n = 1964; 900 μatm, n = 10,818) and total alkalinity (AT; 350 μatm, n = 56; 900 μatm, n = 47) are measured values. All
other values were calculated using the carbonate chemistry system analysis program CO2SYS: Seawater pCO2, dissolved inorganic carbon
(DIC), bicarbonate (HCO3

−), carbonate (CO3
2−), carbon dioxide (CO2), saturation states for calcite (Ωcalcite) and aragonite (Ωaragonite).

Values are presented as mean, with standard deviation below

Location pHT
Salinity
(psu)

AT
(μmol kg−1)

pCO2
(μatm)

DIC
(μmol kg−1)

HCO3
−

(μmol kg−1)
CO3

2−
(μmol kg−1) Ωcalcite Ωaragonite

‘350 μatm’ 8.137 34.504 2264.29 316.057 1962.694 1740.629 211.979 5.087 3.301

0.056 0.427 15.34 47.466 34.376 55.084 22.221 0.534 0.348

‘900 μatm’ 7.788 34.351 2268.33 841.148 2125.785 1984.889 115.150 2.771 1.805

0.106 0.484 19.45 291.762 39.381 52.510 21.308 0.512 0.336

Note: ‘350 μatm’ carbonate chemistry data are sourced from Agostini et al. (2018), ‘900 μatm’ carbonate chemistry data are averaged across Agostini
et al. (2018) and original data collected between 2017/05/31 and 2017/08/08.

    |  5HARVEY Et Al.

community composition and total cover of the sessile fauna (lo-
cated on the underside of the tile) was assessed, following the same
procedure as the community analysis (based on the percent cover
of the sessile faunal species).

2.4  |  Statistical analysis

Statistical analyses were conducted using R (version 3.6.0; R Core
Team, 2019), with the ‘vegan’ (Oksanen et al., 2019), ‘mvabund’
(Wang et al., 2012) and base ‘stats’ package used for statistical anal-
ysis, and the ‘ggplot2’ (Wickham, 2016) and ‘ggpubr’ (Kassambara,
2019) packages used for figure production. For each of the analyses
performed, the package and specific function used in R are listed
below as ‘package::function’.

For the seasonal experiment, differences in community compo-
sition (based on percentage cover) between pCO2 (two levels: 350
and 900 μatm) and Season (two levels: ‘Cold Period’ and ‘Warm
Period’) were assessed using non-metric multidimensional scaling
(nMDS; vegan::metaMDS) and a permutational analysis of variance
(PERMANOVA) based on Bray–Curtis dissimilarity (vegan::vegdist
and vegan::adonis). To test for differences in the percentage cover
of the specific algal functional groups, we employed a two-way
generalized linear model (GLM; Family: Binomial (Link = Logit)),
with pCO2 (two levels: 350 and 900 μatm) and Season (two levels:
‘Cold Period’ and ‘Warm Period’) as fixed factors (stats::glm).

For the reciprocal transplant experiment, differences in commu-
nity composition (based on percentage cover) between Community
Origin (two levels: 350 and 900 μatm), Community Destination (two
levels: 350 and 900 μatm) and Time (two levels: 6 and 12 months) as
fixed factors were assessed using an nMDS (vegan::metaMDS) and
a PERMANOVA based on Bray–Curtis dissimilarity (vegan::vegdist
and vegan::adonis). As the starting communities at 6 months could
influence the resulting communities at 12 months on a particular
tile, we accounted for this repeated measure in the PERMANOVA by
using the ‘strata’ argument within vegan::adonis. To test for differ-
ences in the percentage cover of the specific algal functional groups
at 12 months, we employed a two-way GLM (Family: Binomial
(Link = Logit)), with Community Origin (two levels: 350 and 900 μatm)
and Community Destination (two levels: 350 and 900 μatm) as fixed
factors (stats::glm).

Differences in community production and respiration of the
12 month communities were tested using a two-way GLM (Family:
Gaussian (Link = Identity)), with Community Origin (two levels: 350
and 900 μatm) and Community Destination (two levels: 350 and
900 μatm) as fixed factors (stats::glm). To assess whether changes
in respiration were driven by an altered sessile invertebrate commu-
nity for the 12 month communities, we assessed for differences in
the community composition (based on percentage cover) and total
percentage cover of the sessile fauna. Differences in the community
composition of the sessile fauna, with Community Origin (two lev-
els: 350 and 900 μatm), and Community Destination (two levels: 350
and 900 μatm) as fixed factors, were assessed using a PERMANOVA

based on Bray–Curtis dissimilarity (vegan::vegdist and vegan::adonis).
Differences in the total percentage cover of the sessile fauna were
tested using a two-way GLM (Family: Gaussian (Link = Identity)),
with Community Origin (two levels: 350 and 900 μatm) and
Community Destination (two levels: 350 and 900 μatm) as fixed fac-
tors (stats::glm).

Post hoc comparisons for all PERMANOVAs were achieved
using a Bonferroni-corrected pairwise PERMANOVA. For the re-
ciprocal transplant, only comparisons chosen a priori were tested;
these were between Time (two levels: 6 and 12 months) for each
of the four combinations of Community Origin (two levels: 350
and 900 μatm) and Community Destination (two levels: 350 and
900 μatm). The assumptions of the generalized linear models
(GLMs) were met, with the response variable independent, the
mean–variance relationship suitable (assessed by plotting the
residuals vs. fits; mvabund::manyglm) and dispersion parameters
not being under- or over-dispersed (assessed using quasibinomial
error distributions in R; stats::glm). Post hoc testing of the GLM
for the gross primary production was achieved using TukeyHSD
(multcomp::glht).

3   |   R E S U LT S

3.1  |  Effects of pCO2 and season on community
composition

Overall community composition following 6 months settlement
was highly separated by nMDS, being significantly affected by both
‘pCO2’ and ‘Season’ (PERMANOVA: pCO2*Season, F1,25 = 4.351,
p = 0.012; Figure 2a; Table 2). Community composition was simi-
lar between seasons at 350 μatm (post hoc Bonferroni-adjusted,
p = 0.102), but greatly differed between seasons at 900 μatm (post
hoc Bonferroni-adjusted, p = 0.018). The typical newly settled com-
munity at 350 μatm during the ‘Warm Period’ largely comprised
of a low coverage (<10%) of microalgae and turf algae (Figure 2b) as well as corticated foliose algae, and a high coverage (~75%) of both corticated macrophytes (Figure 2c, including Chondracanthus tenellus (Harvey) Hommersand, 1993) and crustose coralline algae (Figure 2d, including Lithophyllum sp. Philippi, 1837). The community at 350 μatm during the ‘Cold Period’ was similar, the only difference being a significantly lower cover (approximately half) of corticated macrophytes relative to the ‘Warm Period’ (Figure 2c).

Relative to 350 CO2, at 900 μatm the newly settled community
during the ‘Warm Period’ showed increased coverage by microalgae
and turf (from almost zero to ~30% coverage; Figure 2b), as well as
increased corticated foliose algae (tenfold increase; including the
species Zonaria diesingiana J. Agardh, 1841), but a threefold de-
crease in corticated macrophytes (Figure 2c) and a fourfold decrease
in crustose coralline algae (Figure 2d). This resulted in a community
with roughly similar coverage of the microalgae and turf, corticated
macrophytes and crustose coralline algae. During the ‘Cold Period’,
the 900 μatm communities showed a high coverage of microalgae

6  |    HARVEY Et Al.

and turf algae (~72%; Figure 2b), but an absence of corticated foliose
algae. Both the corticated macrophytes (Figure 2c) and crustose cor-
alline algae (Figure 2d) were similar in coverage between the ‘Warm
Period’ and ‘Cold Period’ at 900 μatm CO2.

3.2  |  Effects of early-stage composition and pCO2
on subsequent community succession

After performing a reciprocal transplant of the established
6 month communities for a further 6 months, it was found that

regardless of the conditions under which the communities were
established (i.e. ‘Community Origin’), the communities converged
to form similar communities based on the pCO2 conditions they
were currently residing in (i.e. ‘Community Destination’; Figure 3;
Table 3).

At 12 months the 350 μatm CO2 recruitment tiles typically had
a corticated macrophyte upperstorey (Figure 4b, including Codium
coactum Okamura, 1930 and Chondracanthus tenellus), with an un-
derstorey of microalgae and turf algae (Figure 4a), as well as crustose
coralline algae (Figure 4c, including Lithophyllum sp.). Both the ‘350
to 350 μatm’ and the ‘900 to 350 μatm’ communities had a higher
total cover than at 6 months (significant increase by ~40% cover for
both). This was due to an increased cover of understorey microalgae
and turf algae in the ‘350 to 350 μatm’ community (Figure 4a), and
increased corticated macrophytes (Figure 4b; both as understorey
and canopy) for both communities. Overall the ‘350 to 350 μatm’
community was most complex, followed by the ‘900 to 350 μatm’
community (Figure 4d).

At 900 μatm CO2, a typical community after 12 months lacked
canopy algae and was predominantly comprised of turf algae
(Figure 4a, including Biddulphia biddulphiana (J.E. Smith) Boyer
1900) with some corticated foliose algae (including Zonaria diesing-
iana), and minimal cover of corticated macrophytes (Figure 4b) and

F I G U R E 2  Community composition and percentage cover of algal functional groups. (a) nMDS of community composition based on
algal functional groups. Treatments are displayed by pCO2 (‘350 μatm’ – blue; ‘900 μatm’ – red) and season (‘Warm Period’ – triangles; ‘Cold
Period’ – crosses). (b–d) Percentage cover (%) of (b) microalgae and turf algae, (c) corticated macrophytes, and (d) crustose coralline algae
following 6 months settlement at either 350 μatm pCO2 (‘Warm Period’ – darker blue, ‘Cold Period’ – lighter blue) or 900 μatm pCO2 (‘Warm
Period’ – darker red, ‘Cold Period’ – lighter red). Two-way GLM (pCO2*Season) results are presented in the top-right of (b)–(d). See Table S1
for more detailed statistics

(a)

(c)

T A B L E 2  PERMANOVA summary of pCO2 (350 μatm vs.
900 μatm) and Season (‘Cold Period’ vs. ‘Warm Period’) for the algal
communities. For p-values, * (p < 0.05), ** (p < 0.01), *** (p < 0.001)

Term df
Sum
Sq.

Mean
Sq. F p

pCO2 1 1.995 1.995 32.58 0.001***

Season 1 0.378 0.378 6.18 0.003**

pCO2*Season 1 0.266 0.266 4.35 0.012*

Residuals 22 1.347 0.061

Total 25 3.986

    |  7HARVEY Et Al.

crustose coralline algae (Figure 4c). The community that was trans-
planted into 900 μatm (‘350 to 900 μatm’) showed a decline in total
cover (significant decrease by ~45% cover), and a large reduction in
complexity (Figure 4d) due to the loss of corticated macrophytes
and crustose coralline algae (Figure 4b,c). The community that was

consistently maintained at 900 μatm (‘900 to 900 μatm’) showed an
increased coverage by turf and microalgae (Figure 4a), but overall
displayed similar levels of total cover (~100% cover at 6 months and
~120% cover at 12 months) and complexity (Figure 4d) relative to
the initial 6 month community.

F I G U R E 3  nMDS of community
composition based on algal functional
groups. Communities are grouped as
those exposed to 350 μatm throughout
(darker blue), transplanted from 900 to
350 μatm (lighter blue), transplanted from
350 to 900 μatm (lighter red) and exposed
to 900 μatm throughout (darker red). The
initial starting 6 month communities are
displayed with open symbols and dashed
lines, and the 12 month communities
following the transplant are displayed with
solid symbols and lines

Term df Sum Sq. Mean Sq. F p

Origin 1 0.910 0.910 13.05 0.001***

Destination 1 0.669 0.669 9.59 0.001***

Time 1 1.267 1.267 18.17 0.001***

Origin*Destination 1 0.141 0.141 2.03 0.002**

Origin*Time 1 0.490 0.490 7.03 0.023*

Destination*Time 1 1.16 1.16 16.66 0.003**

Origin*Destination*Time 1 0.074 0.074 1.06 0.541

Residuals 23 1.604 0.070

Total 30 6.136

T A B L E 3  PERMANOVA summary
of Community Origin (‘Origin’: 350 vs.
900 μatm), Community Destination
(‘Destination’: 350 μatm vs. 900 μatm),
and Time (‘6 months’ vs.’12 months’) for
algal communities grown on settlement
tiles off Shikine Island, Japan. For
p-values, * (p < 0.05), ** (p < 0.01), *** (p < 0.001)

F I G U R E 4  Percentage cover (%) of
functional groups following a further
6 month settlement: (a) microalgae and
turf algae, (b) corticated macrophytes,
(c) and crustose coralline algae. (d)
structural complexity of the communities.
Communities are grouped as those
exposed to 350 μatm throughout (darker
blue), transplanted from 900 to 350 μatm
(lighter blue) or 350 to 900 μatm
(lighter red) and exposed to 900 μatm
throughout (darker red). Two-way GLM
results are presented in the top-right
of each. See Table S2 for more detailed
statistics

(a)

(b)

(c) (d)

p = 0.59
p < 0.01

8  |    HARVEY Et Al.

3.3  |  Effects of early-stage composition and pCO2
on community production

Net community production was reduced for the 350 μatm com-
munities compared to the 900 μatm communities (Figure 5a).
This was due to greatly increased community respiration rates
for those communities at 350 μatm (Figure 5b), which resulted
in the ‘350–350 μatm’ communities showing no net community
production, and the ‘900–350 μatm’ communities having a nega-
tive net community production (Figure 5a). Overall, the resulting
communities had similar levels of gross community production
(Figure 5c). A significant interaction was observed for gross com-
munity production (Figure 5c); however, all post hoc comparisons
were non-significant.

3.4  |  Effects of early-stage composition and pCO2
on sessile invertebrate community

To assess whether increased respiration was driven by an altered
sessile invertebrate community for the 12 month communities, the
community composition and total cover of the sessile fauna was
assessed. The sessile invertebrate community composition of the
12 month communities was altered by both ‘Community Origin’ and
‘Community Destination’ (PERMANOVA: F1,14 = 15.07, p < 0.001 and F1,14 = 3.68, p = 0.042, respectively; Figure S2), with a greater cov- erage of ascidians (Didemnidae) and hydrozoans (Leptomedusae) in the elevated pCO2 conditions (‘350–900 μatm’ and ‘900–900 μatm’) and a greater coverage of polychaetes (Serpulidae) in the reference pCO2 conditions (‘350–350 μatm’ and ‘900–350 μatm’; see Figure S3). However, the overall coverage of sessile fauna did not significantly differ between the treatments (GLM: ‘Community Origin’ t = 1.17, p = 0.27 and ‘Community Destination’ t = −0.18, p = 0.86), with no interaction (GLM: ‘Community Origin’*‘Community Destination’ t = −0.07, p = 0.95).

4   |   D I S C U S S I O N

Ocean acidification alters the competitive abilities of algae and so is
expected to change the structure, composition and functioning of
both coastal and open ocean marine habitats (Cornwall et al., 2017;
Hall-Spencer & Harvey, 2019). Observations at volcanic CO2 seeps in
the photic zone have shown profound ecosystem shifts towards sim-
plified non-calcareous communities that are often algal dominated
with lower biodiversity and reduced ecological complexity (Agostini
et al., 2018; Connell et al., 2018; Foo et al., 2018; González-Delgado
& Hernández, 2018; Kroeker, Gambi, et al., 2013). The underlying
processes by which ocean acidification affects the structure of
shallow-water marine communities are not clearly established and
require additional investigation, although significant progress has
been made in recent years (see Kroeker, Gambi, et al., 2013; Kroeker
et al., 2011, 2012; Porzio et al., 2013; Teixidó et al., 2018; Vizzini
et al., 2017). We found that an enriched CO2 environment had a
positive effect on r-selected, fast-growing microalgae and turf algal
species in the early stages of community development. This species-
poor, low complexity, early-successional stage was then locked-in as
it inhibited the settlement and growth of corticated macrophytes.
We highlight the potential ecological processes responsible for this
change in a temperate rocky reef community exposed to enriched
CO2 conditions.

Ocean acidification altered competitive dominance after
6 months with substrata in the acidified conditions being dominated
by microalgae and turf algae, rather than the corticated macrophytes
and crustose coralline algae that dominated in reference pCO2 condi-
tions. This pattern was temporally consistent, showing that regardless
of the growing season ocean acidification truncates the normal suc-
cessional trajectories of communities (Baggini et al., 2014; Kroeker
et al., 2012). The benefits of seawater acidification to opportunistic

F I G U R E 5  (a) Net community production (mmol O2 h
−1 m−2), (b)

Community respiration (mmol O2 h
−1 m−2), and (c) Gross community

production (mmol O2 h
−1 m−2). Communities are grouped as those

exposed to 350 μatm throughout (darker blue), transplanted
between locations (900 to 350 μatm, lighter blue; and 350 to
900 μatm, lighter red), and exposed to 900 μatm throughout
(darker red). See Table S3 for more detailed statistics

(a)
(b)
(c)

p
p
p

p
p
p
p
p
p

    |  9HARVEY Et Al.

species, such as turf algae, over others (including calcareous species)
is well established (Connell et al., 2013, 2018; Cornwall et al., 2017).
Lowered carbonate saturation is a stressor to calcified macroalgae
(Brinkman & Smith, 2015; Enochs et al., 2015; Fabricius et al., 2011;
Kamenos et al., 2016) whilst turf algae and other fast-growing oppor-
tunistic species can use the additional carbon from ocean acidifica-
tion to grow and compete for resources (Harvey et al., 2019; Kroeker
et al., 2012; Porzio et al., 2013), thereby attaining dominance (Connell
et al., 2018). The shift from typical coastal habitat-forming species
(such as corals and kelp forests) to turf algal dominance causes a loss
of structural complexity and associated ecosystem services (O’Brien
& Scheibling, 2018; Rogers et al., 2014).

Before turf algae overgrew the recruitment tiles in acidified con-
ditions (between 6 and 12 months, see Figure 3b), both crustose
coralline algae and corticated macrophytes recruited onto the sub-
strata. This suggests that the divergence in community composition
was not due to limited recruitment or a physiological intolerance to
the acidified conditions, but was driven by altered competitive in-
teractions (Crook et al., 2016; Kroeker et al., 2012) and/or loss of
compensatory processes (Connell et al., 2018; Ghedini et al., 2015).
Although bottom-up control helps stimulate algal growth on coral
and rocky reefs, grazing pressure determines whether turf algae
dominate (Mumby et al., 2006). The grazing pressure of large benthic
invertebrates in acidified sites is thought to be lowered due to phys-
iological impacts (Calosi et al., 2013; O’Donnell et al., 2010) which
cause their size and abundance to be reduced compared to the ref-
erence pCO2 areas (Connell et al., 2018, White Island, New Zealand;
Harvey et al., unpublished, this site and Vulcano, Italy). Similarly, the
observed number of sea urchin feeding halos has also previously
been found to be reduced in the Ischia CO2 seeps (Kroeker, Gambi,
et al., 2013). Fish communities also play a key role in top-down con-
trol and, in our site, communities included more herbivorous fishes
than the surrounding non-acidified areas (Cattano et al., 2020).
Clearly, the increased turf algae supported a greater herbivorous
fish population than in the reference conditions, but those same
herbivores alone are not able to control the increased growth of the
boosted turf algae (also see Baggini et al., 2015; Connell et al., 2018).
Although outside the scope of this study, this may support the no-
tion that the fish are preferentially consuming different algae other
than the turfs and further reinforcing the ecological shift.

After 12 months, assemblages in reference pCO2 conditions
continued to gain species through time and had developed more
structurally complex communities with clearly defined understorey
and canopy species. The assemblages in the elevated pCO2 became
arrested in terms of their successional development due to competi-
tion for space by the turf algae. A similar overgrowth and dominance
by turf algae was observed on recruitment tiles in the Ischia CO2
seep (Kroeker et al., 2012; Porzio et al., 2013). This community de-
velopment in our study resulted in a similar community composition
at 6 and 12 months with only the abundance of the turf algae being
increased. At 12 months, communities on the tiles were visually in-
distinguishable from the surrounding rocky substrata (Figure S4).
Similar declines in macroalgal diversity with increasing pCO2 have

also been demonstrated in Methana, Greece (Baggini et al., 2014).
The simplification of marine ecosystems has been observed across
CO2 seeps (Agostini et al., 2018; Brown et al., 2018; Cigliano et al.,
2010; Fabricius et al., 2011; Kroeker, Gambi, et al., 2013; Vizzini
et al., 2017), with such changes leading to a functional biodiversity
loss in the system (Teixidó et al., 2018). It is likely that such sim-
ple systems are maintained by reinforcing feedback loops (sediment
trapping, changes in physicochemical environment, and recruit-
ment inhibition) that facilitate turf algal dominance. Turf algae can
inhibit successional development by reducing primary substratum
availability (Airoldi, 1998; Connell & Russell, 2010) and by trapping
sediment which alters settlement surface chemistry and reduces the
survival of other recruits (Airoldi, 2003; Gorman & Connell, 2009).
Such dominance by short-lived species, which then locks the system
in place, can lead to decreased stability in the system (Stachowicz
et al., 2007), with implications for the functioning of the system
under future ocean acidification (Teixidó et al., 2018).

In terms of community dynamics, both the reference and ele-
vated pCO2 conditions appeared to overwhelm any ecological re-
sistance that would have otherwise resisted ecosystem change.
This was demonstrated by the established algal communities that
were transplanted from reference to elevated pCO2 conditions con-
verging (in terms of community composition) to almost match the
community formed under elevated pCO2 conditions (and vice versa).
This suggests that acidification-driven ecological shifts to simplified
turf algae communities will occur regardless of the state that the
community is in, and means that the community successional trajec-
tory is not fixed from the initial bare substratum during primary or
secondary succession. The prevention of such shifts by ecosystem
management will require a focused effort on resilience building in
order to mitigate the future degradation of ecosystems (Billé et al.,
2013; Falkenberg et al., 2013). In contrast, the convergence of the
communities transplanted from elevated pCO2 conditions to the ref-
erence pCO2 conditions could mean that recovery from a degraded
state is possible. This would likely be due to sufficient compensatory
processes at our reference pCO2 location, and/or the turf algae los-
ing its competitive edge in the absence of elevated pCO2. Therefore,
a combination of conservation strategy and meaningful reductions
in atmospheric CO2 emissions could achieve substantial recovery
of the abundance, structure and function of shallow coastal marine
ecosystems (Duarte et al., 2020).

Despite possessing highly divergent communities, gross oxygen
production was similar between all of the transplanted tiles. Net ox-
ygen production, however, was positive in the acidified conditions,
but balanced between productivity and respiration for the reference
pCO2 communities due to elevated respiration. Ecosystems that
are more developed and stable will tend towards rates of oxygen
production and respiration being equal, tending to not accumulate
further biomass. Early-stage ecosystems will tend to have a higher
productivity per biomass, but will be lacking in terms of biomass and
species diversity (Cooke, 1967). This further supports the concept
that the algal community developing under elevated pCO2 is ar-
rested into a typical early-stage community dominated by r-selected

10  |    HARVEY Et Al.

species. Previous studies in CO2 seeps have generally focussed on
the primary production or photophysiology of individual species
of algae (e.g. Celis-Plá et al., 2015; Porzio et al., 2018, 2020) with
the aim of assessing their physiological response to ocean acidifica-
tion, rather than the effects on overall community net production
(making comparisons difficult). The sessile invertebrate communi-
ties differed in community composition between the reference and
elevated pCO2 sites, but not in percentage cover, suggesting that
they were not a sizeable contributor toward such large changes in
net oxygen production. Instead, the decreased net oxygen was likely
driven by the greater algal biomass (as well as low surface to vol-
ume ratio) of the more highly structurally complex reference com-
munity compared to the high surface to biomass ratio of the turf
algae. Taken together, this suggests that the greater net production
stimulated by ocean acidification does not translate into enhanced
ecosystem benefits, such as increased community cover, biomass,
biodiversity or structural complexity, as well as an altered sessile in-
vertebrate community.

Natural analogues provide a number of benefits for advancing
our understanding about the responses of shallow-water marine
communities to ocean acidification conditions, but they are not
perfect analogues. Carbonate chemistry at some CO2 seeps can be
highly variable (Rastrick et al., 2018), and areas in close proximity
to CO2 vents can be enriched in some metals and toxins (Vizzini
et al., 2013; Zitoun et al., 2020). It is possible to reduce such con-
founding factors by avoiding toxic areas and only selecting sites a
suitable distance away, since contamination from hydrothermal flu-
ids can be quickly diluted by mixing with seawater (Agostini et al.,
2015; Pichler et al., 2019). The gas being released at our study site
is 98 ± 3 (SD) % CO2, and although concentrations of hydrogen
sulfide are detected at the main vent, they are below detection
limits ~50 m away from the main vents (Agostini et al., 2015) and
the study site used in this study is more than 300 m away from
the main vent. The possibility remains that other trace elements
or heavy metals (as yet unmeasured) may be present either in the
water or bioaccumulated in biota, as has been shown in other CO2
seeps (Mirasole et al., 2020; Mishra et al., 2020; Vizzini et al., 2013;
Zitoun et al., 2020), which could influence the response of marine
organisms to ocean acidification. An additional consideration for
CO2 seeps is that they demonstrate the consequences of future
ocean acidification but in the absence of concurrent ocean warming
(Hughes et al., 2017), and temperatures will mediate the response
of organisms and communities to future ocean acidification. Such
an issue can be addressed by comparing CO2 seep systems under
different thermal regimes and assess the consistency of responses
(Johnson et al., 2012), or by manipulating temperature along CO2
gradients (Alessi et al., 2019). Despite these caveats, the use of CO2
seeps is still invaluable for providing a window into the future state
of organisms, communities and ecosystems to future ocean acidifi-
cation (Rastrick et al., 2018).

In conclusion, ocean acidification can set the course of suc-
cessional development in algal communities, benefitting turf algae
whilst causing reduced algal biomass, diversity and complexity.

Altered carbonate chemistry can enable opportunistic r-selected
species to competitively exclude other species and lock the com-
munity in a species-poor early successional stage. The ecological
process responsible for this shift in community composition was
not simply altering community trajectory during primary succes-
sion, as the same shift occurred in preestablished communities.
This highlights that without reducing atmospheric CO2 emissions
we may increasingly observe the loss of large algal habitats and the
spread of fast-growing, small opportunistic species that can utilize
additional inorganic carbon. By understanding the ecological pro-
cesses responsible for driving shifts in community composition, we
can begin to better assess how communities are likely to be altered
by ocean acidification. Finally, our results show that the recovery of
shallow-water marine communities is possible if meaningful reduc-
tions in CO2 emissions are implemented, as encouraged by the Paris
Agreement.

A C K N O W L E D G E M E N T S
We thank the technical staff at ‘Shimoda Marine Research Center,
University of Tsukuba’ for their assistance aboard RV Tsukuba II and
at the study site, and the Japan Fisheries agencies of Nijima/Shikine
Island (Tokyo prefecture) for their support. This project was heavily
supported by the ‘International Education and Research Laboratory
Program’, University of Tsukuba. This work was also supported by
JSPS KAKENHI Grant Number 17K17622, and we acknowledge
funding support from the Ministry of Environment, Government of
Japan (Suishinhi: 4RF-1701).

Some of the images used within the graphical abstract are courtesy
of the Integration and Application Network, University of Maryland
Center for Environmental Science (ian.umces.edu/symbo ls/).

C O N F L I C T O F I N T E R E S T
The authors declare no conflicts of interest.

A U T H O R C O N T R I B U T I O N S
B.P.H. conceived the idea, designed the methodology, analysed
the data and led the writing of the manuscript. B.P.H. and K.K.
carried out the image analysis. B.P.H. and S.A. performed the
oxygen production measurements. All authors assisted with field
work, contributed critically to the drafts and gave final approval
for publication.

D ATA AVA I L A B I L I T Y S TAT E M E N T
Biological data (Figures 2–5 and Tables 2 and 3) and associated car-
bonate chemistry data (Figure 1; Figure S1; Table 1) are stored on
Zenodo (https://doi.org/10.5281/zenodo.4280018).

O R C I D
Ben P. Harvey https://orcid.org/0000-0002-4971-1634
Koetsu Kon https://orcid.org/0000-0003-1379-0702
Sylvain Agostini https://orcid.org/0000-0001-9040-9296
Shigeki Wada https://orcid.org/0000-0001-6893-7498
Jason M. Hall-Spencer https://orcid.org/0000-0002-6915-2518

http://ian.umces.edu/symbols/

https://doi.org/10.5281/zenodo.4280018

https://orcid.org/0000-0002-4971-1634

https://orcid.org/0000-0002-4971-1634

https://orcid.org/0000-0003-1379-0702

https://orcid.org/0000-0003-1379-0702

https://orcid.org/0000-0001-9040-9296

https://orcid.org/0000-0001-9040-9296

https://orcid.org/0000-0001-6893-7498

https://orcid.org/0000-0001-6893-7498

https://orcid.org/0000-0002-6915-2518

https://orcid.org/0000-0002-6915-2518

    |  11HARVEY Et Al.

R E F E R E N C E S
Abràmoff, M. D., Magalhães, P. J., & Ram, S. J. (2004). Image processing

with ImageJ. Biophotonics International, 11(7), 36–42.
Agostini, S., Harvey, B. P., Wada, S., Kon, K., Milazzo, M., Inaba, K., &

Hall-Spencer, J. M. (2018). Ocean acidification drives community
shifts towards simplified non-calcified habitats in a subtropi-
cal-temperate transition zone. Scientific Reports, 8, 11354. https://
doi.org/10.1038/s4159 8-018-29251 -7

Agostini, S., Wada, S., Kon, K., Omori, A., Kohtsuka, H., Fujimura, H.,
Tsuchiya, Y., Sato, T., Shinagawa, H., Yamada, Y., & Inaba, K. (2015).
Geochemistry of two shallow CO2 seeps in Shikine Island (Japan)
and their potential for ocean acidification research. Regional Studies
in Marine Science, 2(Suppl.), 45–53.

Airoldi, L. (1998). Roles of disturbance, sediment stress, and substratum re-
tention on spatial dominance in algal turf. Ecology, 79(8), 2759–2770.

Airoldi, L. (2003). The effects of sedimentation on rocky coast assem-
blages. Oceanography and Marine Biology: An Annual Review, 41,
169–171.

Albright, R., Caldeira, L., Hosfelt, J., Kwiatkowski, L., Maclaren, J. K.,
Mason, B. M., Nebuchina, Y., Ninokawa, A., Pongratz, J., Ricke, K.
L., Rivlin, T., Schneider, K., Sesboüé, M., Shamberger, K., Silverman,
J., Wolfe, K., Zhu, K., & Caldeira, K. (2016). Reversal of ocean acid-
ification enhances net coral reef calcification. Nature, 531(7594),
362–365.

Albright, R., Takeshita, Y., Koweek, D. A., Ninokawa, A., Wolfe, K.,
Rivlin, T., Nebuchina, Y., Young, J., & Caldeira, K. (2018). Carbon
dioxide addition to coral reef waters suppresses net community
calcification. Nature, 555, 516–519. https://doi.org/10.1038/
natur e25968

Alessi, C., Giomi, F., Furnari, F., Sarà, G., Chemello, R., & Milazzo, M. (2019).
Ocean acidification and elevated temperature negatively affect re-
cruitment, oxygen consumption and calcification of the reef-building
Dendropoma cristatum early life stages: Evidence from a manipula-
tive field study. Science of The Total Environment, 693, 133476.

Algueró-Muñiz, M., Alvarez-Fernandez, S., Thor, P., Bach, L. T., Esposito,
M., Horn, H. G., Ecker, U., Langer, J. A. F., Taucher, J., Malzahn, A.
M., Riebesell, U., & Boersma, M. (2017). Ocean acidification ef-
fects on mesozooplankton community development: Results from
a long-term mesocosm experiment. PLoS One, 12(4), e0175851.
https://doi.org/10.1371/journ al.pone.0175851

Baggini, C., Issaris, Y., Salomidi, M., & Hall-Spencer, J. (2015). Herbivore
diversity improves benthic community resilience to ocean acidi-
fication. Journal of Experimental Marine Biology and Ecology, 469,
98–104. https://doi.org/10.1016/j.jembe.2015.04.019

Baggini, C., Salomidi, M., Voutsinas, E., Bray, L., Krasakopoulou, E., &
Hall-Spencer, J. M. (2014). Seasonality affects macroalgal commu-
nity response to increases in pCO2. PLoS One, 9(9), e106520.

Billé, R., Kelly, R., Biastoch, A., Harrould-Kolieb, E., Herr, D., Joos, F.,
Kroeker, K., Laffoley, D., Oschlies, A., & Gattuso, J.-P. (2013). Taking
action against ocean acidification: A review of management and
policy options. Environmental Management, 52(4), 761–779. https://
doi.org/10.1007/s0026 7-013-0132-7

Brinkman, T. J., & Smith, A. M. (2015). Effect of climate change on crus-
tose coralline algae at a temperate vent site, White Island, New
Zealand. Marine and Freshwater Research, 66(4), 360–370.

Brown, N. E. M., Milazzo, M., Rastrick, S. P. S., Hall-Spencer, J. M.,
Therriault, T. W., & Harley, C. D. G. (2018). Natural acidification
changes the timing and rate of succession, alters community struc-
ture, and increases homogeneity in marine biofouling communities.
Global Change Biology, 24(1), e112–e127. https://doi.org/10.1111/
gcb.13856

Brown, N. E. M., Therriault, T. W., & Harley, C. D. G. (2016). Field-based
experimental acidification alters fouling community structure and
reduces diversity. Journal of Animal Ecology, 85(5), 1328–1339.
https://doi.org/10.1111/1365-2656.12557

Calosi, P., Rastrick, S. P. S., Graziano, M., Thomas, S. C., Baggini, C.,
Carter, H. A., Hall-Spencer, J. M., Milazzo, M., & Spicer, J. I. (2013).
Distribution of sea urchins living near shallow water CO2 vents is
dependent upon species acid-base and ion-regulatory abilities.
Marine Pollution Bulletin, 73(2), 470–484.

Cattano, C., Agostini, S., Harvey, B. P., Wada, S., Quattrocchi, F., Turco,
G., Inaba, K., Hall-Spencer, J. M., & Milazzo, M. (2020). Changes in
fish communities due to benthic habitat shifts under ocean acidi-
fication conditions. Science of The Total Environment, 725, 138501.
https://doi.org/10.1016/j.scito tenv.2020.138501

Celis-Plá, P. S. M., Hall-Spencer, J. M., Horta, P. A., Milazzo, M.,
Korbee, N., Cornwall, C. E., & Figueroa, F. L. (2015). Macroalgal
responses to ocean acidification depend on nutrient and light
levels. Frontiers in Marine Science, 2, 26. https://doi.org/10.3389/
fmars.2015.00026

Cigliano, M., Gambi, M. C., Rodolfo-Metalpa, R., Patti, F. P., & Hall-
Spencer, J. M. (2010). Effects of ocean acidification on invertebrate
settlement at volcanic CO2 vents. Marine Biology, 157(11), 2489–
2502. https://doi.org/10.1007/s0022 7-010-1513-6

Connell, J. H., & Slatyer, R. O. (1977). Mechanisms of succession in nat-
ural communities and their role in community stability and organi-
zation. The American Naturalist, 111(982), 1119–1144. https://doi.
org/10.1086/283241

Connell, S. D., Doubleday, Z. A., Foster, N. R., Hamlyn, S. B., Harley, C.
D., Helmuth, B., Kelaher, B. P., Nagelkerken, I., Rodgers, K. L., Sarà,
G., & Russell, B. D. (2018). The duality of ocean acidification as a
resource and a stressor. Ecology, 99(5), 1005–1010. https://doi.
org/10.1002/ecy.2209

Connell, S. D., Kroeker, K. J., Fabricius, K. E., Kline, D. I., & Russell, B.
D. (2013). The other ocean acidification problem: CO2 as a re-
source among competitors for ecosystem dominance. Philosophical
Transactions of the Royal Society B: Biological Sciences, 368(1627),
20120442. https://doi.org/10.1098/rstb.2012.0442

Connell, S. D., & Russell, B. D. (2010). The direct effects of increasing
CO2 and temperature on non-calcifying organisms: Increasing the
potential for phase shifts in kelp forests. Proceedings of the Royal
Society B: Biological Sciences, 277(1686), 1409–1415.

Cooke, G. D. (1967). The pattern of autotrophic succession in lab-
oratory microcosms. BioScience, 17(10), 717–721. https://doi.
org/10.2307/1294089

Cornwall, C. E., Revill, A. T., Hall-Spencer, J. M., Milazzo, M., Raven, J.
A., & Hurd, C. L. (2017). Inorganic carbon physiology underpins
macroalgal responses to elevated CO2. Scientific Reports, 7, 46297.
https://doi.org/10.1038/srep4 6297

Crook, E. D., Kroeker, K. J., Potts, D. C., Rebolledo-Vieyra, M., Hernandez-
Terrones, L. M., & Paytan, A. (2016). Recruitment and succession in
a tropical benthic community in response to in-situ ocean acidifi-
cation. PLoS One, 11(1), e0146707. https://doi.org/10.1371/journ
al.pone.0146707

Done, T. (1992). Effects of tropical cyclone waves on ecological and geo-
morphological structures on the Great Barrier Reef. Continental
Shelf Research, 12(7–8), 859–872.

Duarte, C. M., Agusti, S., Barbier, E., Britten, G. L., Castilla, J. C., Gattuso,
J.-P., Fulweiler, R. W., Hughes, T. P., Knowlton, N., Lovelock, C. E.,
Lotze, H. K., Predragovic, M., Poloczanska, E., Roberts, C., & Worm,
B. (2020). Rebuilding marine life. Nature, 580(7801), 39–51. https://
doi.org/10.1038/s4158 6-020-2146-7

Enochs, I. C., Manzello, D. P., Donham, E. M., Kolodziej, G., Okano,
R., Johnston, L., Young, C., Iguel, J., Edwards, C. B., Fox, M. D.,
Valentino, L., Johnson, S., Benavente, D., Clark, S. J., Carlton, R.,
Burton, T., Eynaud, Y., & Price, N. N. (2015). Shift from coral to
macroalgae dominance on a volcanically acidified reef. Nature
Climate Change, 5(12), 1083–1088. https://doi.org/10.1038/nclim
ate2758

Fabricius, K. E., Langdon, C., Uthicke, S., Humphrey, C., Noonan, S.,
De’ath, G., Okazaki, R., Muehllehner, N., Glas, M. S., & Lough, J.

https://doi.org/10.1038/s41598-018-29251-7

https://doi.org/10.1038/s41598-018-29251-7

https://doi.org/10.1038/nature25968

https://doi.org/10.1038/nature25968

https://doi.org/10.1371/journal.pone.0175851

https://doi.org/10.1016/j.jembe.2015.04.019

https://doi.org/10.1007/s00267-013-0132-7

https://doi.org/10.1007/s00267-013-0132-7

https://doi.org/10.1111/gcb.13856

https://doi.org/10.1111/gcb.13856

https://doi.org/10.1111/1365-2656.12557

https://doi.org/10.1016/j.scitotenv.2020.138501

https://doi.org/10.3389/fmars.2015.00026

https://doi.org/10.3389/fmars.2015.00026

https://doi.org/10.1007/s00227-010-1513-6

https://doi.org/10.1086/283241

https://doi.org/10.1086/283241

https://doi.org/10.1002/ecy.2209

https://doi.org/10.1002/ecy.2209

https://doi.org/10.1098/rstb.2012.0442

https://doi.org/10.2307/1294089

https://doi.org/10.2307/1294089

https://doi.org/10.1038/srep46297

https://doi.org/10.1371/journal.pone.0146707

https://doi.org/10.1371/journal.pone.0146707

https://doi.org/10.1038/s41586-020-2146-7

https://doi.org/10.1038/s41586-020-2146-7

https://doi.org/10.1038/nclimate2758

https://doi.org/10.1038/nclimate2758

12  |    HARVEY Et Al.

M. (2011). Losers and winners in coral reefs acclimatized to ele-
vated carbon dioxide concentrations. Nature Climate Change, 1(3),
165–169. https://doi.org/10.1038/nclim ate1122

Falkenberg, L. J., Connell, S. D., & Russell, B. D. (2013). Disrupting the
effects of synergies between stressors: Improved water quality
dampens the effects of future CO2 on a marine habitat. Journal of
Applied Ecology, 50(1), 51–58. https://doi.org/10.1111/1365-2664.
12019

Foo, S. A., Byrne, M., Ricevuto, E., & Gambi, M. C. (2018). The carbon
dioxide vents of Ischia, Italy, a natural system to assess impacts
of ocean acidification on marine ecosystems: An overview of re-
search and comparisons with other vent systems. Oceanography
and Marine Biology: An Annual Review, 56, 237–310. https://doi.
org/10.1201/97804 29454 455-9

Garilli, V., Rodolfo-Metalpa, R., Scuderi, D., Brusca, L., Parrinello, D.,
Rastrick, S. P. S., Foggo, A., Twitchett, R. J., Hall-Spencer, J. M., &
Milazzo, M. (2015). Physiological advantages of dwarfing in surviv-
ing extinctions in high-CO2 oceans. Nature Climate Change, 5(7),
678–682. https://doi.org/10.1038/nclim ate2616

Gattuso, J.-P., Magnan, A., Billé, R., Cheung, W. W. L., Howes, E. L.,
Joos, F., Allemand, D., Bopp, L., Cooley, S. R., Eakin, C. M., Hoegh-
Guldberg, O., Kelly, R. P., Pörtner, H.-O., Rogers, A. D., Baxter, J.
M., Laffoley, D., Osborn, D., Rankovic, A., Rochette, J., … Turley,
C. (2015). Contrasting futures for ocean and society from differ-
ent anthropogenic CO2 emissions scenarios. Science, 349(6243),
aac4722. https://doi.org/10.1126/scien ce.aac4722

Gaylord, B., Kroeker, K. J., Sunday, J. M., Anderson, K. M., Barry, J.
P., Brown, N. E., Connell, S. D., Dupont, S., Fabricius, K. E., Hall-
Spencer, J. M., Klinger, T., Milazzo, M., Munday, P. L., Russell, B.
D., Sanford, E., Schreiber, S. J., Thiyagarajan, V., Vaughan, M. L.
H., Widdicombe, S., & Harley, C. D. G. (2015). Ocean acidification
through the lens of ecological theory. Ecology, 96(1), 3–15.

Ghedini, G., & Connell, S. D. (2016). Organismal homeostasis buffers the
effects of abiotic change on community dynamics. Ecology, 97(10),
2671–2679. https://doi.org/10.1002/ecy.1488

Ghedini, G., Russell, B. D., & Connell, S. D. (2015). Trophic compensation
reinforces resistance: Herbivory absorbs the increasing effects of
multiple disturbances. Ecology Letters, 18(2), 182–187. https://doi.
org/10.1111/ele.12405

Godbold, J. A., & Solan, M. (2013). Long-term effects of warming and
ocean acidification are modified by seasonal variation in species re-
sponses and environmental conditions. Philosophical Transactions of
the Royal Society B: Biological Sciences, 368(1627), 20130186.

González-Delgado, S., & Hernández, J. C. (2018). The importance of
natural acidified systems in the study of ocean acidification: What
have we learned? Advances in Marine Biology, 80, 57–99. https://doi.
org/10.1016/bs.amb.2018.08.001

Gordillo, F. J., Figueroa, F. L., & Niell, F. X. (2003). Photon-and carbon-use
efficiency in Ulva rigida at different CO2 and N levels. Planta, 218(2),
315–322.

Gorman, D., & Connell, S. D. (2009). Recovering subtidal forests in hu-
man-dominated landscapes. Journal of Applied Ecology, 46(6), 1258–
1265. https://doi.org/10.1111/j.1365-2664.2009.01711.x

Gruner, D. S., Smith, J. E., Seabloom, E. W., Sandin, S. A., Ngai, J. T.,
Hillebrand, H., Harpole, W. S., Elser, J. J., Cleland, E. E., Bracken, M.
E. S., Borer, E. T., & Bolker, B. M. (2008). A cross-system synthesis
of consumer and nutrient resource control on producer biomass.
Ecology Letters, 11(7), 740–755.

Hall-Spencer, J. M., & Harvey, B. P. (2019). Ocean acidification impacts
on coastal ecosystem services due to habitat degradation. Emerging
Topics in Life Sciences, 3(2), 197–206. https://doi.org/10.1042/
ETLS2 0180117

Hall-Spencer, J. M., Rodolfo-Metalpa, R., Martin, S., Ransome, E., Fine, M.,
Turner, S. M., Rowley, S. J., Tedesco, D., & Buia, M.-C. (2008). Volcanic
carbon dioxide vents show ecosystem effects of ocean acidification.
Nature, 454(7200), 96–99. https://doi.org/10.1038/natur e07051

Harvey, B. P., Agostini, S., Kon, K., Wada, S., & Hall-Spencer, J. M. (2019).
Diatoms dominate and alter marine food-webs when CO2 rises.
Diversity, 11(12). https://doi.org/10.3390/d1112 0242

Harvey, B. P., Agostini, S., Wada, S., Inaba, K., & Hall-Spencer, J. M. (2018).
Dissolution: The Achilles’ heel of the triton shell in an acidifying
ocean. Frontiers in Marine Science, 5, 371. https://doi.org/10.3389/
fmars.2018.00371

Harvey, B. P., Gwynn-Jones, D., & Moore, P. J. (2013). Meta-analysis
reveals complex marine biological responses to the interactive
effects of ocean acidification and warming. Ecology and Evolution,
3(4), 1016–1030. https://doi.org/10.1002/ece3.516

Harvey, B. P., McKeown, N. J., Rastrick, S. P. S., Bertolini, C., Foggo, A.,
Graham, H., Hall-Spencer, J. M., Milazzo, M., Shaw, P. W., Small, D.
P., & Moore, P. J. (2016). Individual and population-level responses
to ocean acidification. Scientific Reports, 6, 20194. https://doi.
org/10.1038/srep2 0194

Hillebrand, H., Gruner, D. S., Borer, E. T., Bracken, M. E. S., Cleland, E.
E., Elser, J. J., Harpole, W. S., Ngai, J. T., Seabloom, E. W., Shurin,
J. B., & Smith, J. E. (2007). Consumer versus resource control of
producer diversity depends on ecosystem type and producer com-
munity structure. Proceedings of the National Academy of Sciences
of the United States of America, 104(26), 10904–10909. https://doi.
org/10.1073/pnas.07019 18104

Hughes, T. P., Barnes, M. L., Bellwood, D. R., Cinner, J. E., Cumming, G.
S., Jackson, J. B. C., Kleypas, J., van de Leemput, I. A., Lough, J. M.,
Morrison, T. H., Palumbi, S. R., van Nes, E. H., & Scheffer, M. (2017).
Coral reefs in the Anthropocene. Nature, 546(7656), 82–90. https://
doi.org/10.1038/natur e22901

Hughes, T. P., Rodrigues, M. J., Bellwood, D. R., Ceccarelli, D., Hoegh-
Guldberg, O., McCook, L., Moltschaniwskyj, N., Pratchett, M. S.,
Steneck, R. S., & Willis, B. (2007). Phase shifts, herbivory, and the
resilience of coral reefs to climate change. Current Biology, 17(4),
360–365. https://doi.org/10.1016/j.cub.2006.12.049

IPCC. (2013). Climate change 2013 – The physical science basis: Working
group I contribution to the fifth assessment report of the IPCC (No.
0521880092; p. 1535). Cambridge University Press.

Jenkins, S. R., Hawkins, S. J., & Norton, T. A. (1999). Interaction between
a fucoid canopy and limpet grazing in structuring a low shore inter-
tidal community. Journal of Experimental Marine Biology and Ecology,
233(1), 41–63.

Johnson, C. R., & Mann, K. H. (1988). Diversity, patterns of adaptation,
and stability of Nova Scotian kelp beds. Ecological Monographs,
58(2), 129–154. https://doi.org/10.2307/1942464

Johnson, V. R., Russell, B. D., Fabricius, K. E., Brownlee, C., & Hall-
Spencer, J. M. (2012). Temperate and tropical brown macroalgae
thrive, despite decalcification, along natural CO2 gradients. Global
Change Biology, 18(9), 2792–2803.

Kamenos, N. A., Perna, G., Gambi, M. C., Micheli, F., & Kroeker, K. J.
(2016). Coralline algae in a naturally acidified ecosystem persist
by maintaining control of skeletal mineralogy and size. Proceedings
of the Royal Society B: Biological Sciences, 283(1840), 20161159.
https://doi.org/10.1098/rspb.2016.1159

Kassambara, A. (2019). ggpubr: “ggplot2” based publication ready plots. R
Package Version 0.2.4. https://CRAN.R-proje ct.org/packa ge=ggpubr

Kelly, E. L. A., Eynaud, Y., Clements, S. M., Gleason, M., Sparks, R. T., Williams,
I. D., & Smith, J. E. (2016). Investigating functional redundancy versus
complementarity in Hawaiian herbivorous coral reef fishes. Oecologia,
182(4), 1151–1163. https://doi.org/10.1007/s0044 2-016-3724-0

Kerfahi, D., Harvey, B. P., Agostini, S., Kon, K., Huang, R., Adams, J. M., &
Hall-Spencer, J. M. (2020). Responses of intertidal bacterial biofilm
communities to increasing pCO2. Marine Biotechnology. https://doi.
org/10.1007/s1012 6-020-09958 -3

Koch, M., Bowes, G., Ross, C., & Zhang, X.-H. (2013). Climate change and
ocean acidification effects on seagrasses and marine macroalgae.
Global Change Biology, 19(1), 103–132. https://doi.org/10.1111/
j.1365-2486.2012.02791.x

https://doi.org/10.1038/nclimate1122

https://doi.org/10.1111/1365-2664.12019

https://doi.org/10.1111/1365-2664.12019

https://doi.org/10.1201/9780429454455-9

https://doi.org/10.1201/9780429454455-9

https://doi.org/10.1038/nclimate2616

https://doi.org/10.1126/science.aac4722

https://doi.org/10.1002/ecy.1488

https://doi.org/10.1111/ele.12405

https://doi.org/10.1111/ele.12405

https://doi.org/10.1016/bs.amb.2018.08.001

https://doi.org/10.1016/bs.amb.2018.08.001

https://doi.org/10.1111/j.1365-2664.2009.01711.x

https://doi.org/10.1042/ETLS20180117

https://doi.org/10.1042/ETLS20180117

https://doi.org/10.1038/nature07051

https://doi.org/10.3390/d11120242

https://doi.org/10.3389/fmars.2018.00371

https://doi.org/10.3389/fmars.2018.00371

https://doi.org/10.1002/ece3.516

https://doi.org/10.1038/srep20194

https://doi.org/10.1038/srep20194

https://doi.org/10.1073/pnas.0701918104

https://doi.org/10.1073/pnas.0701918104

https://doi.org/10.1038/nature22901

https://doi.org/10.1038/nature22901

https://doi.org/10.1016/j.cub.2006.12.049

https://doi.org/10.2307/1942464

https://doi.org/10.1098/rspb.2016.1159

https://CRAN.R-project.org/package=ggpubr

https://doi.org/10.1007/s00442-016-3724-0

https://doi.org/10.1007/s10126-020-09958-3

https://doi.org/10.1007/s10126-020-09958-3

https://doi.org/10.1111/j.1365-2486.2012.02791.x

https://doi.org/10.1111/j.1365-2486.2012.02791.x

    |  13HARVEY Et Al.

Kroeker, K. J., Gambi, M. C., & Micheli, F. (2013). Community dynam-
ics and ecosystem simplification in a high-CO2 ocean. Proceedings
of the National Academy of Sciences of the United States of America,
110(31), 12721–12726. https://doi.org/10.1073/pnas.12164 64110

Kroeker, K. J., Kordas, R. L., Crim, R., Hendriks, I. E., Ramajo, L., Singh, G.
S., Duarte, C. M., & Gattuso, J.-P. (2013). Impacts of ocean acidifica-
tion on marine organisms: Quantifying sensitivities and interaction
with warming. Global Change Biology, 19(6), 1884–1896. https://doi.
org/10.1111/gcb.12179

Kroeker, K. J., Micheli, F., & Gambi, M. C. (2012). Ocean acidification
causes ecosystem shifts via altered competitive interactions.
Nature Climate Change, 3(2), 156–159.

Kroeker, K. J., Micheli, F., Gambi, M. C., & Martz, T. R. (2011). Divergent
ecosystem responses within a benthic marine community to ocean
acidification. Proceedings of the National Academy of Sciences of
the United States of America, 108(35), 14515–14520. https://doi.
org/10.1073/pnas.11077 89108

Li, W., Gao, K., & Beardall, J. (2012). Interactive effects of ocean acid-
ification and nitrogen-limitation on the diatom Phaeodactylum tri-
cornutum. PLoS One, 7(12), e51590. https://doi.org/10.1371/journ
al.pone.0051590

Ling, S. D., Scheibling, R. E., Rassweiler, A., Johnson, C. R., Shears, N.,
Connell, S. D., Salomon, A. K., Norderhaug, K. M., Pérez-Matus, A.,
Hernández, J. C., & Clemente, S. (2015). Global regime shift dynam-
ics of catastrophic sea urchin overgrazing. Philosophical Transactions
of the Royal Society B: Biological Sciences, 370(1659), 20130269.

Milazzo, M., Alessi, C., Quattrocchi, F., Chemello, R., D’Agostaro, R., Gil,
J., Vaccaro, A. M., Mirto, S., Gristina, M., & Badalamenti, F. (2019).
Biogenic habitat shifts under long-term ocean acidification show
nonlinear community responses and unbalanced functions of asso-
ciated invertebrates. Science of The Total Environment, 667, 41–48.
https://doi.org/10.1016/j.scito tenv.2019.02.391

Milazzo, M., Rodolfo-Metalpa, R., Chan, V. B. S., Fine, M., Alessi, C.,
Thiyagarajan, V., Hall-Spencer, J. M., & Chemello, R. (2014). Ocean
acidification impairs vermetid reef recruitment. Scientific Reports, 4,
4189. https://doi.org/10.1038/srep0 4189

Mirasole, A., Scopelliti, G., Tramati, C., Signa, G., Mazzola, A., & Vizzini,
S. (2020). Evidences on alterations in skeleton composition and
mineralization in a site-attached fish under naturally acidified
conditions in a shallow CO2 vent. Science of The Total Environment,
143309. https://doi.org/10.1016/j.scito tenv.2020.143309

Mishra, A. K., Santos, R., & Hall -Spencer, J. M. (2020). Elevated trace
elements in sediments and seagrasses at CO2 seeps. Marine
Environmental Research, 153, 104810. https://doi.org/10.1016/j.
maren vres.2019.104810

Moulin, L., Grosjean, P., Leblud, J., Batigny, A., Collard, M., & Dubois, P.
(2015). Long-term mesocosms study of the effects of ocean acid-
ification on growth and physiology of the sea urchin Echinometra
mathaei. Marine Environmental Research, 103, 103–114. https://doi.
org/10.1016/j.maren vres.2014.11.009

Mumby, P. J., Dahlgren, C. P., Harborne, A. R., Kappel, C. V., Micheli,
F., Brumbaugh, D. R., Holmes, K. E., Mendes, J. M., Broad, K., &
Sanchirico, J. N. (2006). Fishing, trophic cascades, and the process
of grazing on coral reefs. Science, 311(5757), 98–101.

O’Brien, J., & Scheibling, R. (2018). Turf wars: Competition between foun-
dation and turf-forming species on temperate and tropical reefs and
its role in regime shifts. Marine Ecology Progress Series, 590, 1–17.

O’Donnell, M. J., Todgham, A. E., Sewell, M. A., Hammond, L. M.,
Ruggiero, K., Fangue, N. A., Zippay, M. L., & Hofmann, G. E. (2010).
Ocean acidification alters skeletogenesis and gene expression in
larval sea urchins. Marine Ecology Progress Series, 398, 157–171.

Oksanen, J., Blanchet, F. G., Friendly, M., Kindt, R., Legendre, P., McGlinn,
D., Minchin, P. R., O’Hara, R. B., Simpson, G. L., Solymos, P., Stevens,
M. H. H., Szoecs, E., & Wagner, H. (2019). The vegan package.
Vegan: Community ecology package. R Package Version 2.5-5. https://
Cran.r-Proje ct.Org/Packa ge=vegan

Pichler, T., Biscéré, T., Kinch, J., Zampighi, M., Houlbrèque, F., & Rodolfo-
Metalpa, R. (2019). Suitability of the shallow water hydrothermal
system at Ambitle Island (Papua New Guinea) to study the effect
of high pCO2 on coral reefs. Marine Pollution Bulletin, 138, 148–158.
https://doi.org/10.1016/j.marpo lbul.2018.11.003

Porzio, L., Arena, C., Lorenti, M., De Maio, A., & Buia, M. C. (2020).
Long-term response of Dictyota dichotoma var. intricata (C. Agardh)
Greville (Phaeophyceae) to ocean acidification: Insights from high
pCO2 vents. Science of The Total Environment, 731, 138896. https://
doi.org/10.1016/j.scito tenv.2020.138896

Porzio, L., Buia, M. C., Ferretti, V., Lorenti, M., Rossi, M., Trifuoggi,
M., Vergara, A., & Arena, C. (2018). Photosynthesis and mineral-
ogy of Jania rubens at low pH/high pCO2: A future perspective.
Science of The Total Environment, 628–629, 375–383. https://doi.
org/10.1016/j.scito tenv.2018.02.065

Porzio, L., Garrard, S. L., & Buia, M. C. (2013). The effect of ocean acid-
ification on early algal colonization stages at natural CO2 vents.
Marine Biology, 160(8), 2247–2259. https://doi.org/10.1007/s0022
7-013-2251-3

R Core Team. (2019). R: A language and environment for statistical comput-
ing. R Foundation for Statistical Computing. Retrieved from https://
www.R-proje ct.org/

Rastrick, S. S. P., Graham, H., Azetsu-Scott, K., Calosi, P., Chierici, M.,
Fransson, A., Hop, H., Hall-Spencer, J., Milazzo, M., Thor, P., &
Kutti, T. (2018). Using natural analogues to investigate the effects
of climate change and ocean acidification on Northern ecosys-
tems. ICES Journal of Marine Science, 75(7), 2299–2311. https://doi.
org/10.1093/icesj ms/fsy128

Rogers, A., Blanchard, J. L., & Mumby, P. J. (2014). Vulnerability of coral
reef fisheries to a loss of structural complexity. Current Biology,
24(9), 1000–1005. https://doi.org/10.1016/j.cub.2014.03.026

Stachowicz, J. J., Bruno, J. F., & Duffy, J. E. (2007). Understanding
the effects of marine biodiversity on communities and ecosys-
tems. Annual Review of Ecology Evolution and Systematics, 38,
739–766.

Steneck, R. S., & Dethier, M. N. (1994). A functional group approach
to the structure of algal-dominated communities. Oikos, 69(3),
476–498.

Sunday, J. M., Fabricius, K. E., Kroeker, K. J., Anderson, K. M., Brown, N.
E., Barry, J. P., Connell, S. D., Dupont, S., Gaylord, B., Hall-Spencer,
J. M., Klinger, T., Milazzo, M., Munday, P. L., Russell, B. D., Sanford,
E., Thiyagarajan, V., Vaughan, M. L. H., Widdicombe, S., & Harley, C.
D. G. (2017). Ocean acidification can mediate biodiversity shifts by
changing biogenic habitat. Nature Climate Change, 7, 81–85. https://
doi.org/10.1038/nclim ate3161

Teixidó, N., Gambi, M. C., Parravacini, V., Kroeker, K., Micheli, F., Villéger,
S., & Ballesteros, E. (2018). Functional biodiversity loss along nat-
ural CO2 gradients. Nature Communications, 9(1), 5149. https://doi.
org/10.1038/s4146 7-018-07592 -1

Vizzini, S., Di Leonardo, R., Costa, V., Tramati, C. D., Luzzu, F., &
Mazzola, A. (2013). Trace element bias in the use of CO2 vents as
analogues for low pH environments: Implications for contamina-
tion levels in acidified oceans. Estuarine, Coastal and Shelf Science,
134, 19–30.

Vizzini, S., Martínez-Crego, B., Andolina, C., Massa-Gallucci, A., Connell,
S. D., & Gambi, M. C. (2017). Ocean acidification as a driver of com-
munity simplification via the collapse of higher-order and rise of
lower-order consumers. Scientific Reports, 7(1), 4018. https://doi.
org/10.1038/s4159 8-017-03802 -w

Wang, Y., Naumann, U., Wright, S. T., & Warton, D. I. (2012). Mvabund–
An R package for model-based analysis of multivariate abundance
data. Methods in Ecology and Evolution, 3(3), 471–474.

Wickham, H. (2016). ggplot2: Elegant graphics for data analysis.
Springer-Verlag.

Wilson, S. S., Furman, B. T., Hall, M. O., & Fourqurean, J. W. (2020). Assessment
of hurricane Irma impacts on south Florida seagrass communities using

https://doi.org/10.1073/pnas.1216464110

https://doi.org/10.1111/gcb.12179

https://doi.org/10.1111/gcb.12179

https://doi.org/10.1073/pnas.1107789108

https://doi.org/10.1073/pnas.1107789108

https://doi.org/10.1371/journal.pone.0051590

https://doi.org/10.1371/journal.pone.0051590

https://doi.org/10.1016/j.scitotenv.2019.02.391

https://doi.org/10.1038/srep04189

https://doi.org/10.1016/j.scitotenv.2020.143309

https://doi.org/10.1016/j.marenvres.2019.104810

https://doi.org/10.1016/j.marenvres.2019.104810

https://doi.org/10.1016/j.marenvres.2014.11.009

https://doi.org/10.1016/j.marenvres.2014.11.009

https://Cran.r-Project.Org/Package=vegan

https://Cran.r-Project.Org/Package=vegan

https://doi.org/10.1016/j.marpolbul.2018.11.003

https://doi.org/10.1016/j.scitotenv.2020.138896

https://doi.org/10.1016/j.scitotenv.2020.138896

https://doi.org/10.1016/j.scitotenv.2018.02.065

https://doi.org/10.1016/j.scitotenv.2018.02.065

https://doi.org/10.1007/s00227-013-2251-3

https://doi.org/10.1007/s00227-013-2251-3

https://www.R-project.org/

https://www.R-project.org/

https://doi.org/10.1093/icesjms/fsy128

https://doi.org/10.1093/icesjms/fsy128

https://doi.org/10.1016/j.cub.2014.03.026

https://doi.org/10.1038/nclimate3161

https://doi.org/10.1038/nclimate3161

https://doi.org/10.1038/s41467-018-07592-1

https://doi.org/10.1038/s41467-018-07592-1

https://doi.org/10.1038/s41598-017-03802-w

https://doi.org/10.1038/s41598-017-03802-w

14  |    HARVEY Et Al.

long-term monitoring programs. Estuaries and Coasts, 43(5), 1119–
1132. https://doi.org/10.1007/s1223 7-019-00623 -0

Witkowski, C. R., Agostini, S., Harvey, B. P., van der Meer, M. T. J.,
Sinninghe Damsté, J. S., & Schouten, S. (2019). Validation of car-
bon isotope fractionation in algal lipids as a $p$CO2 proxy using
a natural CO2 seep (Shikine Island, Japan). Biogeosciences, 16(22),
4451–4461. https://doi.org/10.5194/bg-16-4451-2019

Zitoun, R., Connell, S. D., Cornwall, C. E., Currie, K. I., Fabricius, K.,
Hoffmann, L. J., Lamare, M. D., Murdoch, J., Noonan, S., Sander,
S. G., Sewell, M. A., Shears, N. T., van den Berg, C. M. G., & Smith,
A. M. (2020). A unique temperate rocky coastal hydrothermal vent
system (Whakaari–White Island, Bay of Plenty, New Zealand):
Constraints for ocean acidification studies. Marine and Freshwater
Research, 71(3), 321–344.

S U P P O R T I N G I N F O R M AT I O N
Additional supporting information may be found online in the
Supporting Information section.

How to cite this article: Harvey BP, Kon K, Agostini S, Wada
S, Hall-Spencer JM. Ocean acidification locks algal
communities in a species-poor early successional stage. Glob
Change Biol. 2020;00:1–14. https://doi.org/10.1111/
gcb.15455

https://doi.org/10.1007/s12237-019-00623-0

https://doi.org/10.5194/bg-16-4451-2019

https://doi.org/10.1111/gcb.15455

https://doi.org/10.1111/gcb.15455

L E T T E R
Ocean acidification reduces coral recruitment by disrupting

intimate larval-algal settlement interactions

Christopher Doropoulos,1,2*

Selina Ward,1 Guillermo

Diaz-Pulido,2,3 Ove

Hoegh-Guldberg2,4 and

Peter J. Mumby1,2

Abstract
Successful recruitment in shallow reef ecosystems often involves specific cues that connect planktonic

invertebrate larvae with particular crustose coralline algae (CCA) during settlement. While ocean acidification

(OA) can reduce larval settlement and the abundance of CCA, the impact of OA on the interactions between

planktonic larvae and their preferred settlement substrate are unknown. Here, we demonstrate that CO2
concentrations (800 and 1300 latm) predicted to occur by the end of this century significantly reduce coral
(Acropora millepora) settlement and CCA cover by ‡ 45%. The CCA important for inducing coral settlement
(Titanoderma spp., Hydrolithon spp.) were the most deleteriously affected by OA. Surprisingly, the only preferred

settlement substrate (Titanoderma) in the experimental controls was avoided by coral larvae as pCO2 increased,

and other substrata selected. Our results suggest OA may reduce coral population recovery by reducing coral

settlement rates, disrupting larval settlement behaviour, and reducing the availability of the most desirable

coralline algal species for successful coral recruitment.

Keywords
Acropora, coral, crustose coralline algae, electivity, Hydrolithon, ocean acidification, recruitment, settlement,

Titanoderma.

Ecology Letters (2012) 15: 338–346

INTRODUCTION

The effects of ocean acidification (OA) have raised concerns about

coral reef ecosystem function by reducing the calcification rates of

benthic organisms important to maintaining habitat structure and

biodiversity (Hoegh-Guldberg et al. 2007; Kroeker et al. 2010).

Anthropogenic emissions of carbon dioxide (CO2) have increased

atmospheric CO2 from approximately 280 ppm prior to the year 1750

to > 380 ppm in 2005 (Jansen et al. 2007), and these are continuing to

rise (Le Quere et al. 2009). The absorption of this atmospheric CO2 by

the oceans has reduced global pH by 0.1 units and carbonate

saturation state by 20% since 1800 (Orr et al. 2005). Numerous

laboratory studies have demonstrated that corals (Schneider & Erez

2006; Anthony et al. 2008), calcifying algae (Anthony et al. 2008;

Kuffner et al. 2008), and coral reef communities (Langdon et al. 2000;

Andersson et al. 2009) have reduced calcification in seawater with

lower pH due to depleted carbonate saturation.

Ecological processes pivotal to coral reef resilience, including coral

recruitment, herbivory, trophic integrity, and connectivity (Knowlton

2001; Mumby et al. 2007), under high CO2 levels have hardly been

investigated (Doney et al. 2009). Yet, growing evidence suggests that

interactions between species are altered as CO2 increases. Under

conditions of OA, corals in contact with fleshy macroalgae had higher

mortality (Diaz-Pulido et al. 2011), and fish mortality increased as OA

reduced the ability of juvenile fish to detect their predators (Munday

et al. 2010). Furthermore, it has been suggested that turf algae can

decrease the recruitment of crustose coralline algae (CCA) (Kuffner

et al. 2008; Russell et al. 2009) and kelp (Connell & Russell 2010)

because of greater space occupation at elevated pCO2. While these

examples illustrate that ecological interactions can be altered as CO2
increases, potential interactions of OA on coral recruitment have not

been addressed.

Recruitment is critical to community recovery as it represents a

crucial process in the development of populations in the post-

disturbance period. A key ecological process in the formation of

coral reefs is the settlement of coral larvae from the plankton to

the reef substrata. Many larvae test benthic substrates for

microhabitat suitability prior to settlement (i.e. attachment and

metamorphosis), with the selection of optimal microhabitats critical

in the post-settlement survival of benthic invertebrates (Raimondi

& Keough 1990; Harrington et al. 2004). Different benthic algae

offer both inductive and inhibitive settlement cues for planktonic

invertebrate larvae (Rodriguez et al. 1993; Kuffner et al. 2006; Diaz-

Pulido et al. 2010), and larvae often search for appropriate substrata

associated with specific CCA and microbial communities for

successful settlement (Morse et al. 1988; Johnson & Sutton 1994;

Heyward & Negri 1999; Negri et al. 2001; Webster et al. 2004).

While recent evidence demonstrates the settlement of coral larvae is

reduced as pCO2 increases (Albright et al. 2010; Albright &

Langdon 2011; Nakamura et al. 2011), the interactions between

planktonic larvae and the CCA community under elevated CO2
levels are unknown.

1School of Biological Sciences, University of Queensland, St Lucia, Qld 4072,

Australia
2Australian Research Council Centre of Excellence for Coral Reef Studies,

University of Queensland, St Lucia, Qld 4072, Australia

3Griffith School of Environment and Australian Rivers Institute, Nathan Campus,

Griffith University, Nathan, QLD 4111, Australia
4Global Change Institute, University of Queensland, St Lucia, Qld 4072, Australia

*Correspondence: E-mail: c.doropoulos@uq.edu.au

Ecology Letters, (2012) 15: 338–346 doi: 10.1111/j.1461-0248.2012.01743.x

� 2012 Blackwell Publishing Ltd/CNRS

Here, we test the hypothesis that elevated pCO2 (400 control, 800

and 1300 latm) alters the recruitment of a spawning coral (Acropora
millepora) by affecting the benthic algal community structure, and the

interactions between the substrata and larvae during settlement. We

used a mechanistic approach with three complementary experiments

to investigate how OA reduces larval settlement. First, to investigate

whether OA caused a shift in the community structure of the

settlement substrata to alter coral settlement, we preconditioned

settlement tiles in treatment seawater for 60 days prior to conducting

6 day settlement assays on those tiles in ambient seawater (expt. 1).

Second, we conducted the reciprocal experiment by isolating the

exposure of elevated pCO2 seawater to the coral larvae and settlement

substrata during the 6 days settlement phase only (expt. 2). Finally, we

explored whether there was a combined effect on coral settlement

when the settlement substrata and coral larvae were both exposed to

elevated pCO2 for 60 and 6 days respectively (expt. 3). From this

series of experiments, we show that OA decreases coral settlement

rates by reducing the availability of specific CCA preferred for larval

settlement, as well as interfering with the interaction between larvae

and CCA by altering the settlement behaviour of the coral larvae, such

that previously avoided substrata are preferentially selected as pCO2
increases in all three conditions.

MATERIAL AND METHODS

CO2 treatments and general protocol

Coral settlement experiments were conducted from October to

December 2009, at Heron Island Research Station, southern Great

Barrier Reef (GBR). Settlement substrata and coral larvae were

exposed to three treatments, which represented control (pH 8.04,

401 latm), and two elevated (pH 7.79, 807 latm; pH 7.60,
1299 latm) levels of future CO2 concentrations (Table 1). Treatments
were based on the worst-case stabilisation levels V (pCO2 700–

850 latm) and VI (pCO2 > 900 latm) predicted by the Intergovern-
mental Panel on Climate Change (IPCC) (Meehl et al. 2007). These

were chosen for the experiment as current CO2 emissions are tracking

the most carbon intensive levels (A1FI) predicted by the IPCC (Le

Quere et al. 2009).

As pH is reduced in a predictable manner by elevated pCO2, the

CO2 levels of the experimental seawater were controlled by adjusting

the pH of the seawater in 200 L sumps (Table 1) (see Diaz-Pulido

et al. 2011 for system details). Briefly, the total pH of the seawater was

continuously measured with temperature compensated pH electrodes

(InPro4501VP; Mettler-Toledo, Melbourne, Victoria, Australia),

which maintained the targeted pH levels with a control unit

(Aquatronica, AEB technologies, Italy) that opened solenoid valves

that injected CO2 into the seawater when pH exceeded the desired

threshold. The calibration of the pH probes was checked daily, and

recalibrated with Mettler-Toledo calibration buffers to 0.01 pH units

when necessary. Alkalinity was measured on seawater samples taken

every 6 h over a spring tidal cycle (2.4 m range) at the end of the study

period to capture the largest variation in the seawater alkalinity and

consolidate the pH treatments to the CO2 levels. Alkalinity replicates

within a sample were analysed until a maximum 2% error was met,

using a Metrohm auto-titrator at Edith Cowan University, WA. The

carbonate chemistry of the control and experimental seawater was

calculated with CO2SYS (Lewis & Wallace 2006) using pH, total

alkalinity, salinity (35.4 ppt ± 0.2 SEM; n = 8), and temperature as

the inputs, with the constants from Mehrbach et al. (1973) refitted by

Dickson & Millero (1987).

The settlement tile CO2 conditioning and settlement assays were

conducted on tiles in replicate tanks in the outdoor flow-through

aquarium system (details of the CO2 exposure times, tank and tile

replication are described below in the protocols under each

experiment and in supplementary Fig. S1). The three treatments were

fed from the 200 L sumps into replicate 12 L tanks at a mean flow

rate of 2.4 (± 0.2 SEM) L min
)1

, and each tank had a small

powerhead for extra seawater circulation. This flow rate and water

movement maintained the target pH levels, which were verified

regularly with a portable SG2 SevenGo
TM

pH meter. Replicate tanks

were randomised on the aquarium table under shade-cloth to account

for the heterogeneity in light, which averaged 406 (± 18 SEM)

lmol m)2 s)1 between 6 AM to 6 PM.

Settlement tile preparation

Unglazed terracotta settlement tiles (� 5 · 5 · 0.5 cm) were initially
preconditioned on the Heron Island reef flat (23� 26¢ 42.2¢¢ S, 151�
54¢ 47.0¢¢ E) for 5 months to develop a microbial and encrusting
community important to coral settlement (Heyward & Negri 1999).

Tiles were collected and carefully cleaned of fouling organisms using a

toothbrush, tweezers, and a plastic scraper. The tiles were then

randomly placed in replicate 12 L aquaria and conditioned in the

control and elevated CO2 treatments for 60 days prior to the

settlement assays. During this time the walls of the aquaria were

cleaned regularly to minimize any algal growth. Settlement tiles were

orientated horizontally at the bottom of the tanks and were stacked in

tile pairs with a 0.5 cm spacer, maximising the amount of cryptic

surfaces available for settlement, as coral larvae generally settle in

cryptic areas in shallow habitats (Wallace 1985).

Coral larvae collection

Gravid adult colonies of Acropora millepora were located on the Heron

Island reef flat around the time of the predicted spawning (2nd Dec

2009). Acropora millepora was chosen as a model organism as it is

Table 1 Summary of the physical and chemical seawater values for CO2 treatment levels

Treatment

Temp*

pH*

TA* pCO2� HCO3
)� CO3

2)�

XAragonite��C lmol kg)1 latm lmol kg)1 lmol kg)1

Control 26.0 (± 0.6) 8.04 (± 0.01) 2355 (± 14) 401 (± 11) 1800 (± 20) 227 (± 4) 3.6 (± 0.07)

Stabilisation level V 26.0 (± 0.6) 7.79 (± 0.01) 2365 (± 20) 807 (± 14) 2019 (± 23) 142 (± 3) 2.3 (± 0.06)

Stabilisation level VI 26.0 (± 0.6) 7.60 (± 0.01) 2363 (± 19) 1299 (± 21) 2125 (± 19) 97 (± 3) 1.6 (± 0.06)

*Temperature, pH, and total alkalinity are means (± SEM) of five replicates.

�pCO2, bicarbonate, carbonate and aragonite saturation state (X) were calculated using CO2 SYS (Lewis & Wallace 2006).

Letter Elevated CO2 alters CCA-larval interactions 339

� 2012 Blackwell Publishing Ltd/CNRS

commonly found on GBR and Indo-Pacific shallow reef flats. Five

colonies were collected and transported to outdoor aquarium facilities

where they were housed in 60 L flow-through aquaria until they

released their egg-sperm bundles. The bundles were broken apart by

gently stirring and agitating the water, and gametes from the different

colonies were collected and cross-fertilized. Fertilization took place

for 2 h, after which the embryos were collected and reared in a

laboratory at 25 �C in ambient seawater, using 200 L sumps with
aeration. At least half the seawater was changed every few hours for

the first 24 h and every 6–12 h thereafter. This removed dead larvae

and unfertilized gametes, to minimise contamination of the rearing

sumps. The larvae developed cilia and began swimming 3 days after

spawning, after which they were used for the settlement assays.

Swimming A. millepora larvae were randomly removed from the

rearing sumps, added to the experimental aquaria, and allowed 6 days

to settle (i.e. attach and metamorphose). The number of larvae added

to each tank during the settlement assays was standardised to 150

(± 10) per tile. After this time, the tiles were removed from the tanks

at random and inspected for settlement with a dissecting microscope.

Experiment 1

To isolate whether changes to the benthic community altered coral

settlement, settlement tiles were conditioned at 400, 800, and

1300 latm pCO2 for 60 days. Following the conditioning period,
coral settlement assays were conducted for 6 days on those tiles with

control seawater only. Three replicate tanks per treatment, with 8 tiles

and 1200 (± 80) larvae per tank, were used for

the experiment (Fig. S1).

Experiment 2

A reciprocal experiment was conducted to determine whether

settlement was altered by elevated pCO2 exposure of the coral larvae

and the benthic community during the settlement assays only.

Settlement assays were conducted for 6 days using the three CO2
seawater treatments described above with settlement tiles that were

conditioned with control seawater only. Two replicate tanks per

treatment, with 6 tiles and 900 (± 60) larvae per tank, were used for

the experiment (Fig. S1).

Experiment 3

Finally, to investigate the combined effect of prolonged exposure of

elevated pCO2 on the benthic community and the settling larvae,

settlement tiles were conditioned in the three CO2 treatments for

60 days prior to conducting 6 day settlement assays on those tiles in

the treatment seawater described above. Three replicate tanks per

treatment, with 10 tiles and 1500 (± 150) larvae per tank, were used

for the experiment (Fig. S1).

Response variables and data analyses

We analysed total larval settlement, benthic community and CCA

community cover of the settlement tiles, and coral settlement

substrate preferences for each of the three experiments. The number

of settled (i.e. attached and metamorphosed) coral larvae was initially

quantified for all orientations of each tile. However, we only analysed

the undersides of each tile (for this and all other variables) as the

number of corals settled in this orientation accounted for ‡ 95% of
the total settlement.

The benthic community of the settlement tiles was quantified by

placing a grid on a tile, and evaluating the dominant substrate in a

square (7.5 mm
2
) using a dissecting microscope, with 224–377 squares

per tile. The substrata were characterised into eight major benthic

groups which were: bare tile, CCA, dead crustose coralline algae

(DCCA), endolithic algae found in dead crustose coralline algae

(EDCCA), turf algae found on dead crustose coralline algae

(TDCCA), turf algae (Turf), encrusting fleshy algae (EFA), and other

organisms which included biofilm, bryozoans, foraminifera, and other

encrusting organisms (Other). CCA specimens were identified to the

finest taxonomic resolution where possible and included nine CCA

taxa (see Appendix S1 in Supporting Information for details on CCA

identification). When CCA specimens could not be identified to genus

or species, they were placed in to an Unknown CCA group, which

represented � 6% of the total CCA community. See supplementary
Fig. S2 for images of the dominant benthic groups and CCA taxa.

The substrate settled on by each individual was quantified to

investigate larval settlement behaviour using Vanderploeg and Scavia�s
electivity index (E*). This index is analogous to Ivlev�s E, but
incorporates a selectivity coefficient and the number of substrata

available for settlement (Lechowicz 1982). Therefore: E* = [Wa )
(1 ⁄ n)] ⁄ [Wa + (1 ⁄ n)], where n is the total number of substrate types
available on each tile and W is the selectivity coefficient for substrate

�a� determined by: Wa = [ra ⁄ pa] ⁄
P

(ra ⁄ pa),(rb ⁄ pb)…(rz ⁄ pz), r is the
proportion of coral larvae settled on substrata a to z on each tile, and

p is the proportion of substrata a to z available for settlement on each

tile. A substrate was selected at random for larval settlement when E*

was � 0, preferably settled on when E* was > 0, and avoided for
settlement when E* was < 0.

The number of coral larvae settled per tile was analysed with a

generalised linear mixed effects model using Poisson distribution. We

tested the effects of elevated pCO2 on counts of coral settlement

amongst CO2 treatment (3 levels, fixed) with replicate tanks as a

random effect and nested in CO2 treatment. The effect of elevated

pCO2 on the percent cover of the broad benthic community and CCA

community composition were tested using a mixed effects permuta-

tional MANOVA (PERMANOVA), with the same fixed and random

factors described for the previous model. When significant differences

were detected (P < 0.05), pair-wise comparisons were performed to

investigate treatment effects. In multivariate analyses, SIMPER

analysis was used to determine the variables that characterised the

dissimilarity amongst treatments. Univariate ANOVA was conducted

within CCA cover to determine any significant treatment effects. All

percentage cover data were sin
)1 �x transformed to meet require-

ments of homogeneity (permDISP) prior to analysis. Finally, we tested

the effect of CO2 treatment (3 levels, fixed) on coral settlement

behaviour with tanks as replicates using PERMANOVA. In all

PERMANOVA main effect and pair-wise tests, we used the P-values

generated by 99 999 permutations when the number of unique

permutations were large, and the Monte Carlo asymptotic P-value

otherwise (Anderson 2005).

RESULTS

We report the results of each of the three settlement experiments in

turn, describing the impacts of OA on overall coral settlement density,

340 C. Doropoulos et al. Letter

� 2012 Blackwell Publishing Ltd/CNRS

the structure of benthic substrata, and settlement behaviour of the

larvae. Results are summarised in Table 2.

Experiment 1

A reduction in the cover of CCA and shift in the CCA community

from elevated CO2 decreased coral settlement in the OA treatments.

The reduction in settlement decreased significantly from an average of

11.0 individuals per 25 cm
2

in the control, to 1.6 and 5.5 individuals at

800 and 1300 latm, respectively (Table 2; Fig. 1a). The cover of
CCAs changed dramatically in the elevated CO2 treatments with a

significant decline of � 50% (ANOVA: F2,6 = 13.283; P = 0.014;
Table 2; supplementary Fig. S3a). The CCA community structure also

changed as pCO2 increased, with three out of ten coralline algal taxa

declining with increasing CO2 concentrations (MANOVA: F2,6 = 3.286;

P = 0.017; Table 2). Titanoderma spp., Hydrolithon boreale, and

H. farinosum were the species that characterised the loss of CCA

cover in both the elevated CO2 treatments (supplementary Fig. S3b).

The settlement behaviour of the larvae, as measured by their

substrate selectivity, was significantly affected by the exposure of

settlement tiles to elevated pCO2 prior to the settlement assays

(MANOVA: F2,6 = 4.291; P = 0.004; Table 2). Titanoderma spp. was the

only preferred settlement substrate in the control treatment

(E* = 0.8) and there were lower rates of settlement on all other

substrata than would be expected by chance (supplementary Fig. S4a).

At 800 latm, larvae did not show any clear settlement preferences and
most substrata were avoided (supplementary Fig. S4b), while the

larvae showed a weak preference for H. farinosum (E* = 0.2) at

1300 latm (supplementary Fig. S4c).

Experiment 2

As expected, there were no differences between the broad community

composition, the CCA percent cover, or the CCA community

amongst the settlement tiles that were allocated for use in these

settlement assays (Table 2). Yet, exposure of coral larvae and the

settlement tiles to elevated pCO2 during the 6 day settlement assays

significantly reduced coral settlement, as it declined from an average

of 11.9 individuals per 25 cm
2

in the control, to 4.7 and 2.8 individuals

at 800 and 1300 latm, respectively (Table 2; Fig. 1b). A similar
disruption to larval settlement behaviour occurred to that found when

only the tiles were pre-exposed to elevated pCO2 for a prolonged

period of time (exp. 1). Again, coral larvae preferred to settle on

Titanoderma spp. (E* = 0.75) in controls (supplementary Fig. S5a),

most substrata were avoided at 800 latm (supplementary Fig. S5b),
and a weak preference for H. farinosum (E* = 0.2) was found at

1300 latm (supplementary Fig. S5c).

Experiment 3

Again, settlement was reduced when the tiles were conditioned in the

CO2 treatments for 60 days, and 6 day settlement assays were

conducted on those tiles under elevated pCO2. The magnitude of the

effect was similar to whether the tiles were conditioned in the CO2
treatments for 60 days prior to the 6 day settlement assays with

control seawater only (exp.1), or whether the larvae and tiles were

exposed to the CO2 treatments during the 6 day settlement assays

only (exp. 2) (Table 2). Increased CO2 reduced the settlement of A.

millepora from an average of 9.7 individuals per 25 cm
2

in the control,

to 5.2 and 4.2 individuals at 800 and 1300 latm, respectively (Fig. 1c).
The reduction in settlement was significant between the control and

highest CO2 treatment (P = 0.046) and marginally significant between

the control and intermediate treatment (P = 0.060).

The changes in tile community structure were similar to those in

experiment 1, but the effects of OA appeared to be less variable in this

experiment. As a result, the wider benthic community structure on the

tile undersides was found to differ significantly amongst the CO2
treatments (MANOVA: F2,6 = 2.612; P = 0.003; Table 2; Fig. 2a).

The loss of coralline algae was partly replaced by an increase of 8% in

the cover of �bare tile� (Fig. 2a). As in experiment 1, OA led to a
significant reduction in the cover of CCAs on the tiles (ANOVA:

F2,6 = 40.538; P = 0.002; Table 2), characterised by declines in

Titanoderma spp., H. boreale, and H. farinosum (Fig. 2b).

Coral settlement behaviour was again altered significantly by

elevated pCO2 (MANOVA: F2,6 = 4.224; P = 0.004; Table 2;

Fig. 3). Of the 19 substrata available, Titanoderma spp. was again the

only preferred settlement substrate in the control (E* = 0.6), while all

other substrata were avoided (Fig. 3a). At 800 latm, Hydrolithon
reinboldii was the only preferred coral settlement substrate (E* = 0.3),

and all other settlement substrata were either randomly settled on or

avoided (Fig 3b). No substrate was preferred for settlement at

1300 latm, with random settlement on bare tile (E* = )0.05), and all
other substrata were avoided (Fig 3c).

DISCUSSION

In our study, the settlement density of coral larvae decreased by

‡ 45% as pCO2 increased from 400 to 800 and 1300 latm in all three

Table 2 Changes to the response variables in Experiments 1, 2, and 3, comparing elevated CO2 treatments (800 and 1300 latm) to the controls (400 latm)

Response variable

Experiment 1 Experiment 2 Experiment 3

800 latm 1300 latm 800 latm 1300 latm 800 latm 1300 latm

1. Total settlement fl 82%*** fl 45% fl 58%*** fl 75%*** fl 50% fl 60%*
2. Benthic community structure NS NS NS NS CCA CCA**

3. CCA cover fl 47%* fl 52%* NS NS fl 42%* fl 63%***
4. CCA community structure Titanoderma** Titanoderma* NS NS NS Titanoderma**

5a). Overall settlement behaviour Titanoderma* Titanoderma* NS NS Sporolithon* Titanoderma**

5b). Selectivity from Titanoderma fl 72% fl 74% fl 69% fl 65% fl 35% fl60%

SIMPER analysis determined the variable that characterised the difference between the control and elevated CO2 treatments in multivariate analyses. Coral behaviour (5) is

divided into the change in settlement preferences of the larvae (5a) among the substrate community, and (5b) from Titanoderma spp., the only preferred settlement substrate in

the controls. Significance values are indicated by: NS = non-significant, * = < 0.05, ** = < 0.01, *** = < 0.001.

Letter Elevated CO2 alters CCA-larval interactions 341

� 2012 Blackwell Publishing Ltd/CNRS

experiments. The reduction in settlement was accompanied by a

profound decline in the cover of CCA when the settlement substrata

were conditioned in elevated CO2 treatments for 60 days prior to the

settlement assays (expt. 1 & 3). While recent studies have also found

inverse relationships between elevated pCO2 and rates of coral

settlement (Albright et al. 2010; Albright & Langdon 2011; Nakamura

et al. 2011), and overall CCA cover (Hall-Spencer et al. 2008; Kuffner

et al. 2008; Russell et al. 2009; Fabricius et al. 2011), our study is the

first to directly link benthic community cover with coral settlement

and it provides three important novel insights. First, we identified the

most susceptible CCA to OA and found that they are the most

important taxa for coral settlement, particularly Titanoderma. Secondly,

we discovered that OA reduced the affinity between the settling larvae

and Titanoderma, their preferred settlement substrate. Third, we found

that similar changes in settlement behaviour occurred under all three

(a)

(b)

(c)

Figure 1 Coral (Acropora millepora) settlement rates on experimental tiles (25 cm
2
) in

response to increasing pCO2. Assays occurred on (a) settlement tiles conditioned in

treatment seawater for 60 days prior to 6 day larval settlement assays on those tiles

with control seawater (n = 3); (b) settlement tiles and larvae exposed to treatment

seawater for 6 days during the settlement assays on tiles conditioned in control

seawater only (n = 2); and (c) settlement tiles and larvae exposed to treatment

seawater for 60 and 6 days, respectively (n = 3). Data are means ± SEM.

Significance values comparing elevated CO2 treatments to the control are indicated

by: * = < 0.05, ** = < 0.01, *** = < 0.001.

(a)
(b)

Figure 2 Percent cover of (a) the broad benthic community and (b) the crustose

coralline algae community in response to increasing pCO2. Settlement tiles were

exposed to the treatments for 66 days, which involved a 60 day pre-exposure

period prior to the 6 day settlement assays (expt. 3). CCA = crustose coralline

algae. DCCA = dead crustose coralline algae. EDCCA = endolithic algae in dead

crustose coralline algae. TDCCA = turf on dead crustose coralline algae.

Turf = filamentous algal turf. EFA = encrusting fleshy algae. Other = biofilm,

carbonate, bryozoans, encrusting foraminifera, and unidentified. H. boreale = Hy-

drolithon boreale. H. farinosum = Hydrolithon farinosum. H. reinboldii = Hydrolithon

reinboldii. Data are

means ± SEM; n = 3.

342 C. Doropoulos et al. Letter

� 2012 Blackwell Publishing Ltd/CNRS

experimental conditions. As we explain below, this surprising result

implies that coral settlement behaviour is mediated by cues associated

with coralline algae that appear to be highly sensitive to elevated

pCO2.

We designed our experiments to distinguish the effects of OA on

the settling organisms (corals) from the settlement surfaces (the

benthic community on the tiles). In experiment 1, we subjected tiles to

a 60 day exposure to elevated pCO2 that resulted in profound changes

to the coralline algal assemblage. When these tiles were then placed in

control (ambient) conditions with coral planulae, the settlement

behaviour of the larvae was disrupted. Because the larvae never

experienced OA conditions in this experiment, the result implies that

prolonged exposure of substrates to OA may alter the cues associated

with CCA that are used by larvae to settle preferentially on Titanoderma.

To examine the influence of OA on the settling larvae themselves

(expt. 2), we exposed them to OA treatments during the 6 day

settlement assays. In this case, all benthic substrata were precondi-

tioned in control seawater prior to the experiment and were exposed

to the treatment seawater for the 6 day period during the assays.

Again, we found the same qualitative disruption to larval settlement

behaviour, suggesting that the 6 day exposure of the benthic

community to OA disrupted the signalling from the CCA, and

potentially that the larvae may also be directly affected by elevated

pCO2. In the third experiment, we found a similar qualitative result

when larvae were subjected to OA and offered settlement substrates

that had also been exposed to the treatments for a 60 day period.

There are two possible explanations of our results. The most

parsimonious explanation is that even a daily exposure of the benthos

to OA disrupts the signalling mechanisms used by coral planulae to

preferentially settle upon Titanoderma. That is, the outcome for

settlement was the same whether the tiles were pre-conditioned for

60 days, causing profound changes in coralline cover, or 6 days during

the experiment. This explanation is consistent with all three

experiments and accounts for the lack of an additive impact of OA

on settlement when both larvae and substrates were exposed to OA.

The most likely mechanism is that larval settlement behaviour is

mediated by bacteria and ⁄ or chemical cues associated with CCA
because settlement was disrupted even when the cover of coralline

algae was unchanged (exp. 2). These results imply that the cues

associated with algal morphogens and ⁄ or the bacterial communities
associated with the CCA thalli are highly sensitive to changes in water

chemistry. It was recently demonstrated that microbial communities

associated with biofilms grown on glass slides were altered after

11 days in elevated pCO2 (Witt et al. 2011). There is a precedent for

the role of bacteria in facilitating settlement (Johnson & Sutton 1994;

Negri et al. 2001; Webster et al. 2004), but neither the taxon specificity

(to Titanoderma) nor the sensitivity to OA have been shown, and future

work should isolate whether it is changes in bacterial communities

and ⁄ or the morphogens associated with CCA that alters the
preference of larval settlement under elevated CO2.

An alternative, albeit not mutually exclusive, explanation is that

coral settlement on CCA is disrupted by exposure of either partner to

(a)
(b)
(c)

Figure 3 Coral (Acropora millepora) settlement behaviour of the

substrata that the coral larvae preferred (> 0), avoided (< 0), or

randomly (� 0) settled on in response to (a) 400, (b) 800, and (c)
1300 latm pCO2 using Vanderploeg and Scavia�s electivity index
(E*). Settlement assays occurred with settlement tiles and larvae

exposed to the treatments for 60 and 6 days, respectively (expt. 3).

See Fig. 2 for abbreviated substrate definitions. Data are

means ± SEM; n = 3.

Letter Elevated CO2 alters CCA-larval interactions 343

� 2012 Blackwell Publishing Ltd/CNRS

OA conditions (i.e. exposure of either the larvae or the algae). This

explanation is consistent with recent reports of the impacts of elevated

pCO2 on coral larvae metabolic rate (Albright & Langdon 2011;

Nakamura et al. 2011) and metamorphosis (Albright et al. 2010;

Albright & Langdon 2011; Nakamura et al. 2011), and on fish larvae

olfactory ability (Dixson et al. 2010; Munday et al. 2010). It has also

been shown that invertebrate larvae become less discriminating in the

selection of their preferred substrate for settlement when they are

under stress (Marshall & Keough 2003). Thus, the coral larvae may

have lost their selectivity for Titanoderma due to the stress related to

OA. Yet, this explanation is not entirely satisfactory for the following

reasons. Firstly, we have to accept that the similarity in outcome from

manipulating the settlement substrata versus the planulae is coinci-

dental (i.e. the disruption to either partner has the same overall

outcome). Secondly, we cannot easily account for the absence of a

clear additive effect when both partners were perturbed simulta-

neously.

Previous studies of coral recruitment in both spawning and

brooding corals, including those from the families Acroporidae,

Agariciidae, Pocilloporidae, and Poritidae, and stemming from both

the Atlantic and Indo-Pacific, have found that coral larvae have an

innate ability to settle preferentially on a single CCA genus,

Titanoderma, and that ensuing survival is greatest on this substrate

compared to any other (Harrington et al. 2004; Arnold et al. 2010;

Price 2010; Ritson-Williams et al. 2010). Here, we found that OA

presents two problems for settling corals. Not only is Titanoderma

exceptionally sensitive to OA, such that its availability is compro-

mised, but larval behaviour switches from a high preference to settle

on Titanoderma to avoidance. Corals have previously been shown to

settle on other substrata, including Hydrolithon spp. and bare tile that

the larvae preferentially settled upon at the elevated CO2 treatments in

this study, but this occurs at lower rates of settlement and survival

(Harrington et al. 2004; Arnold et al. 2010; Price 2010; Ritson-Williams

et al. 2010). Titanoderma has been proposed to be a good facilitator of

coral settlement because it does not slough off tissue and therefore

provides a persistent substratum for recruits, while some species of

Hydrolithon and other CCA slough their tissue to remove fouling

organisms (Harrington et al. 2004; Ritson-Williams et al. 2010). Thus,

while the settlement of corals onto previously avoided substrata at

elevated pCO2 in our study implies that their post-settlement survival

may be reduced, empirical investigations of the long-term survival of

recruits on different substrates at elevated pCO2 are needed to directly

test this hypothesis as it may be an adaptive trait.

Titanoderma is a cryptic, early successional species with relatively

rapid growth, creeping morphology and delicate, thin thalli

(< 500 lm) (Steneck 1986; Ringeltaube & Harvey 2000; Littler & Littler 2003). Its morphology and cryptic, opportunistic nature make it

indicative of fresh substratum with relatively benign levels of stress

such as parrotfish grazing or sediment scour. Such environments are

likely to be ideal for coral settlement because new substratum is likely

to possess fewer competitors (Vermeij & Sandin 2008) and parrotfish

predation can be problematic for coral recruits (Penin et al. 2010).

However, we hypothesize that some of the traits that make Titanoderma

such an important settlement inducer might predispose a particular

sensitivity to OA. It has previously been demonstrated that elevated

pCO2 decreases the abundance and recruitment of coralline algae, in

both field (Hall-Spencer et al. 2008; Fabricius et al. 2011) and

laboratory (Kuffner et al. 2008; Russell et al. 2009) settings. While

these reports (Hall-Spencer et al. 2008; Kuffner et al. 2008) found an

inverse competitive relationship between CCA and turf cover as pCO2
increased, we found that reduced CCA was accompanied by an

increase in the amount of bare tile rather than turf. This suggests that

the reduction of CCA cover was not a consequence of space

competition with turfs, but a direct effect of OA on CCA. In our

study, the three most sensitive taxa of CCA to elevated pCO2
(H. boreale, H. farinosum, and Titanoderma spp.), are all early successional

species with rapid growth and thin thalli (< 500 lm) (Steneck 1986; Ringeltaube & Harvey 2000; Littler & Littler 2003). In contrast, later

successional CCA taxa that have thicker crusts (> 500 lm) (e.g.
Sporolithon, Neogoniolithon, Porolithon) (Steneck 1986; Ringeltaube &

Harvey 2000; Littler & Littler 2003) may be more resistant to OA.

While further studies are needed to test this hypothesis, our results

show that increasing levels of dissolved CO2 are likely to have

profound consequences for the functional diversity of coralline algal

communities.

Our research has demonstrated the ecological mechanics of how

ocean acidification may interfere with a critical process important to

the resilience of a diverse marine ecosystem. This occurred by a

reduction to the abundance of the preferred substrate for larval

settlement, and a disruption to an intimate ecological interaction

between the coral larvae and its preferred substrate during settlement.

The altered interaction between coral settlement and the CCA

community suggests that future recruitment of individuals may be

impaired by CO2 concentrations predicted to be reached this century.

These type of impacts of increased CO2 on non-trophic ecological

interactions between species are just starting to be experimentally

demonstrated (e.g. Connell & Russell 2010; Dixson et al. 2010; Diaz-

Pulido et al. 2011), but suggest profound consequences on the

recovery potential of shallow marine ecosystems (e.g. coral reefs)

following local and global disturbances.

ACKNOWLEDGEMENTS

We thank M. Cowlin, A. Noel, M. Nitschke, M. Smith, O. McIntosh,

and the staff at Heron Island Research Station for their technical

assistance in the field; K. Anthony for designing and providing the

experimental aquarium system; J. Pandolfi for providing laboratory

facilities to process the tiles; and, to the four anonymous referees who

provided constructive criticism of the original manuscript. This

research was financially supported by an ARC Discovery Grant

awarded to O. Hoegh-Guldberg, S. Ward and G. Diaz-Pulido, a QLD

Smart Futures PhD Scholarship to C. Doropoulos, and an ARC

Laureate Fellowship to PJ. Mumby. All work was conducted under

GBRMPA permit number 31597.1.

AUTHOR CONTRIBUTIONS

CD, SW, GDP and PJM designed the study, CD and SW conducted

the study, CD and GDP collected the data, and CD and PJM analysed

the data. CD wrote the first draft of the manuscript, and all the

authors contributed substantially to the interpretation and final

version of the paper.

REFERENCES

Albright, R. & Langdon, C. (2011). Ocean acidification impacts multiple early life

history processes of the Caribbean coral Porites astreoides. Glob. Change Biol., 17,

2478–2487.

344 C. Doropoulos et al. Letter

� 2012 Blackwell Publishing Ltd/CNRS

Albright, R., Mason, B., Miller, M. & Langdon, C. (2010). Ocean acidification

compromises recruitment success of the threatened Caribbean coral Acropora

palmata. Proc. Natl Acad. Sci. USA, 107, 20400–20404.

Anderson, M.J. (2005). PERMANOVA: A FORTRAN Computer Program for

Permutational Analysis of Variance. Department of Statistics, University of

Auckland, New Zealand.

Andersson, A.J., Kuffner, I.B., Mackenzie, F.T., Jokiel, P.L., Rodgers, K.S. &

Tan, A. (2009). Net loss of CaCO3 from a subtropical calcifying community due

to seawater acidification: mesocosm-scale experimental evidence. Biogeosciences, 6,

1811–1823.

Anthony, K.R.N., Kline, D.I., Diaz-Pulido, G., Dove, S. & Hoegh-Guldberg, O.

(2008). Ocean acidification causes bleaching and productivity loss in coral reef

builders. Proc. Natl Acad. Sci. USA, 105, 17442–17446.

Arnold, S.N., Steneck, R.S. & Mumby, P.J. (2010). Running the gauntlet: inhibitory

effects of algal turfs on the processes of coral recruitment. Mar. Ecol. Prog. Ser.,

414, 91–105.

Connell, S.D. & Russell, B.D. (2010). The direct effects of increasing CO2 and

temperature on non-calcifying organisms: increasing the potential for phase shifts

in kelp forests. P. R. Soc. B-Biol. Sci., 277, 1409–1415.

Diaz-Pulido, G., Harii, S., McCook, L.J. & Hoegh-Guldberg, O. (2010). The impact

of benthic algae on the settlement of a reef-building coral. Coral Reefs, 29, 203–

208.

Diaz-Pulido, G., Gouezo, M., Tilbrook, B., Dove, S. & Anthony, K.R.N. (2011).

High CO2 enhances the competitive strength of seaweeds over corals. Ecol. Lett.,

14, 156–162.

Dickson, A.G. & Millero, F.J. (1987). A comparison of the equilibrium constants

for the dissociation of carbonic acid in seawater media. Deep-Sea Res., 34, 1733–

1743.

Dixson, D.L., Munday, P.L. & Jones, G.P. (2010). Ocean acidification disrupts the

innate ability of fish to detect predator olfactory cues. Ecol. Lett., 13, 68–75.

Doney, S.C., Fabry, V.J., Feely, R.A. & Kleypas, J.A. (2009). Ocean acidification: the

other CO2 problem. Annu. Rev. Mar. Sci., 1, 169–192.

Fabricius, K.E., Langdon, C., Uthicke, S., Humphrey, C., Noonan, S., De�ath, G.
et al. (2011). Losers and winners in coral reefs acclimatized to elevated carbon

dioxide concentrations. Nature Clim. Change, 1, 165–169.

Hall-Spencer, J.M., Rodolfo-Metalpa, R., Martin, S., Ransome, E., Fine, M.,

Turner, S.M. et al. (2008). Volcanic carbon dioxide vents show ecosystem effects

of ocean acidification. Nature, 454, 96–99.

Harrington, L., Fabricius, K., De�Ath, G. & Negri, A. (2004). Recognition and
selection of settlement substrata determine post-settlement survival in corals.

Ecology, 85, 3428–3437.

Heyward, A.J. & Negri, A.P. (1999). Natural inducers for coral larval metamor-

phosis. Coral Reefs, 18, 273–279.

Hoegh-Guldberg, O., Mumby, P.J., Hooten, A.J., Steneck, R.S., Greenfield, P.,

Gomez, E. et al. (2007). Coral reefs under rapid climate change and ocean

acidification. Science, 318, 1737–1742.

Jansen, E., Overpeck, J., Briffa, K.R., Duplessy, J.-C., Joos, F., Masson-Delmotte, V.

et al. (2007). Palaeoclimate. In: Climate Change 2007: The Physical Science Basis.

Contribution of Working Group I to the Fourth Assessment Report of the Intergovernmental

Panel on Climate Change (eds Solomon, S., Qin, D., Manning, M., Chen, Z.,

Marquis, M. & Averyt, K.B., et al.). Cambridge University Press, Cambridge, pp.

433–498.

Johnson, C.R. & Sutton, D.C. (1994). Bacteria on the surface of crustose coralline

algae induce metamorphosis of the crown-of-thorns starfish Acanthaster planci.

Mar. Biol., 120, 305–310.

Knowlton, N. (2001). The future of coral reefs. Proc. Natl Acad. Sci. USA, 98, 5419–

5425.

Kroeker, K.J., Kordas, R.L., Crim, R.N. & Singh, G.G. (2010). Meta-analysis reveals

negative yet variable effects of ocean acidification on marine organisms. Ecol.

Lett., 13, 1419–1434.

Kuffner, I.B., Walters, L.J., Becerro, M.A., Paul, V.J., Ritson-Williams, R. &

Beach, K.S. (2006). Inhibition of coral recruitment by macroalgae and cyano-

bacteria. Mar. Ecol. Prog. Ser., 323, 107–117.

Kuffner, I.B., Andersson, A.J., Jokiel, P.L., Rodgers, K.S. & Mackenzie, F.T. (2008).

Decreased abundance of crustose coralline algae due to ocean acidification.

Nature Geosci., 1, 114–117.

Langdon, C., Takahashi, T., Sweeney, C., Chipman, D., Goddard, J., Marubini, F.

et al. (2000). Effect of calcium carbonate saturation state on the calcification rate

of an experimental coral reef. Global Biogeochem. Cy., 14, 639–654.

Le Quere, C., Raupach, M.R., Canadell, J.G., Marland, G., Bopp, L., Ciais, P. et al.

(2009). Trends in the sources and sinks of carbon dioxide. Nature Geosci., 2, 831–836.

Lechowicz, M.J. (1982). The sampling characteristics of electivity indexes. Oecologia,

52, 22–30.

Lewis, P.D.E. & Wallace, D.W.R. (2006). MS excel program developed for CO2
system calculations. Carbon Dioxide Information Analysis Center, Oak Ridge

National Laboratory, U.S. Department of Energy, Oak Ridge.

Littler, D.S. & Littler, M.M. (2003). South Pacific Reef Plants: A Diver�s Guide to the
Plant Life of South Pacific Coral Reefs. Offshore Graphics, Washington, D.C.

Marshall, D.J. & Keough, M.J. (2003). Variation in the dispersal potential of non-

feeding invertebrate larvae: the desperate larva hypothesis and larval size.

Mar. Ecol. Prog. Ser., 255, 145–153.

Meehl, G.A., Stocker, T.F., Collins, W.D., Friedlingstein, P., Gaye, A.T.,

Gregory, J.M. et al. (2007). Global climate projections. In: Climate Change 2007:

The Physical Science Basis. Contribution of Working Group I to the Fourth Assessment

Report of the Intergovernmental Panel on Climate Change (eds Solomon, S., Qin, D.,

Manning, M., Chen, Z., Marquis, M. & Averyt, K.B., et al.). Cambridge University

Press, Cambridge, pp. 747–845.

Mehrbach, C., Culberso, C.H., Hawley, J.E. & Pytkowic, R.M. (1973). Measurement

of apparent dissociation-constants of carbonic-acid in seawater at atmospheric

pressure. Limnol. Oceanogr., 18, 897–907.

Morse, D.E., Hooker, N., Morse, A.N.C. & Jensen, R.A. (1988). Control of larval

metamorphosis and recruitment in sympatric Agaricid corals. J. Exp. Mar. Biol.

Ecol., 116, 193–217.

Mumby, P.J., Harborne, A.R., Williams, J., Kappel, C.V., Brumbaugh, D.R., Micheli,

F. et al. (2007). Trophic cascade facilitates coral recruitment in a marine reserve.

Proc. Natl Acad. Sci. USA, 104, 8362–8367.

Munday, P.L., Dixson, D.L., McCormick, M.I., Meekan, M., Ferrari, M.C.O. &

Chivers, D.P. (2010). Replenishment of fish populations is threatened by ocean

acidification. Proc. Natl Acad. Sci. USA, 107, 12930–12934.

Nakamura, M., Ohki, S., Suzuki, A. & Sakai, K. (2011). Coral larvae under ocean

acidification: survival, metabolism, and metamorphosis. PLoS ONE, 6, e14521.

Negri, A.P., Webster, N.S., Hill, R.T. & Heyward, A.J. (2001). Metamorphosis of

broadcast spawning corals in response to bacteria isolated from crustose algae.

Mar. Ecol. Prog. Ser., 223, 121–131.

Orr, J.C., Fabry, V.J., Aumont, O., Bopp, L., Doney, S.C., Feely, R.A. et al. (2005).

Anthropogenic ocean acidification over the twenty-first century and its impact on

calcifying organisms. Nature, 437, 681–686.

Penin, L., Michonneau, F., Baird, A.H., Connolly, S.R., Pratchett, M.S., Kayal, M.

et al. (2010). Early post-settlement mortality and the structure of coral assem-

blages. Mar. Ecol. Prog. Ser., 408, 55–64.

Price, N. (2010). Habitat selection, facilitation, and biotic settlement cues affect

distribution and performance of coral recruits in French Polynesia. Oecologia, 163,

747–758.

Raimondi, P.T. & Keough, M.J. (1990). Behavioural variability in marine larvae.

Aust. J. Ecol., 15, 427–437.

Ringeltaube, P. & Harvey, A. (2000). Non-geniculate coralline algae (Corallinales,

Rhodophyta) on Heron Reef, Great Barrier Reef (Australia). Bot. Mar., 43, 431–454.

Ritson-Williams, R., Paul, V.J., Arnold, S.N. & Steneck, R.S. (2010). Larval settle-

ment preferences and post-settlement survival of the threatened Caribbean corals

Acropora palmata and A. cervicornis. Coral Reefs, 29, 71–81.

Rodriguez, S.R., Ojeda, F.P. & Inestrosa, N.C. (1993). Settlement of benthic marine

invertebrates. Mar. Ecol. Prog. Ser., 97, 193–207.

Russell, B.D., Thompson, J.A.I., Falkenberg, L.J. & Connell, S.D. (2009). Syner-

gistic effects of climate change and local stressors: CO2 and nutrient-driven

change in subtidal rocky habitats. Glob. Change Biol., 15, 2153–2162.

Schneider, K. & Erez, J. (2006). The effect of carbonate chemistry on calcification

and photosynthesis in the hermatypic coral Acropora eurystoma. Limnol. Oceanogr.,

51, 1284–1293.

Steneck, R.S. (1986). The ecology of coralline algal crusts: convergent patterns and

adaptive strategies. Annu. Rev. Ecol. Syst., 17, 273–303.

Vermeij, M.J.A. & Sandin, S.A. (2008). Density-dependent settlement and mortality

structure the earliest life phases of a coral population. Ecology, 89, 1994–2004.

Letter Elevated CO2 alters CCA-larval interactions 345

� 2012 Blackwell Publishing Ltd/CNRS

Wallace, C.C. (1985). Seasonal peaks and annual fluctuations in recruitment of

juvenile scleractinian corals. Mar. Ecol. Prog. Ser., 21, 289–298.

Webster, N.S., Smith, L.D., Heyward, A.J., Watts, J.E.M., Webb, R.I., Blackall, L.L.

et al. (2004). Metamorphosis of a scleractinian coral in response to microbial

biofilms. Appl. Environ. Microbiol., 70, 1213–1221.

Witt, V., Wild, C., Anthony, K.R.N., Diaz-Pulido, G. & Uthicke, S. (2011). Effects

of ocean acidification on microbial community composition of, and oxygen

fluxes through, biofilms from the Great Barrier Reef. Environ. Microbiol., 13,

2976–2989.

SUPPORTING INFORMATION

Additional Supporting Information may be downloaded via the online

version of this article at Wiley Online Library (www.ecologyletters.com).

As a service to our authors and readers, this journal provides

supporting information supplied by the authors. Such materials are

peer-reviewed and may be re-organized for online delivery, but are not

copy-edited or typeset. Technical support issues arising from

supporting information (other than missing files) should be addressed

to the authors.

Editor, J. Emmett Duffy

Manuscript received 15 November 2011

First decision made 24 December 2011

Manuscript accepted 8 January 2012

346 C. Doropoulos et al. Letter

� 2012 Blackwell Publishing Ltd/CNRS

Your Name

Interactionsbetween Plant Semiochemicals and Insects

There are many methods of communication prevalent in species interactions. However,
some methods allow species from even different kingdoms to communicate with each other.
Plants, in order to communicate with insects, release signals known as semiochemicals, which
are packets of chemicals used to deliver some sort of message. These semiochemicals vary in
their effects, and different plants have evolved different kinds of semiochemicals for certain
situations. Examining these chemicals allows humans to understand the varying kinds of insect-
plant interactions, as well as give humans a means by which insects can be communicated to
through artificial semiochemicals. These examples of the powerful and efficient effects of
semiochemicals can show us the importance of cross-species communication.

I. There are several introductory elements to semiochemicals which must be known.

Plants interact with insects by the way of ‘odor plumes’ carrying plant volatiles that affect the
insect’s olfactory senses (Beyaert and Hilker, 2014).

Semiochemials, in a general sense, operate with a type of ‘push-pull strategy’ when used on
insects (Cook et al. 2006).

One basic function of plant semiochemicals is to repel insects. Sometimes this has the added
effect of ‘inhibiting’ the insect’s ability to sense pheromones (Reddy and Guerrero, 2004).

Normally, plants release defensive compounds only during the day. Some plants, such as the
tobacco plant, have evolved to release compounds at night to deal nocturnal herbivores (De
Moraes et al. 2001).

II. While some plants use semiochemicals as a basic ‘push’ protection, others are able to
use them to ‘pull’ insects towards them.

Some plants, such as orchids, trick insects into thinking they are potential mates not only through
visual mimicry, but also through semiochemicals (Dettner and Liepert, 1994).

Some kinds of insects take in plant compounds to use as pheromones of their own (Landolt and
Phillips, 1997)

Some plants are able to attract other organisms that prey on the herbivores they are being
attacked by using volatiles (Bernasconi et al. 1998).

There is a specific method by which hunter and parasitoid organisms find their host with plant
semiochemicals (Stowe et al. 1995).

III. Humans have been able to synthesize semiochemicals for their own uses.

Your Name

Humans are able to use semiochemicals as an alternative to normal means of pest control
(Agelopoulos et al. 1999).

There are multiple benefits for using semiochemicals for pest management compared to
insecticides (Witzgall et al. 2010).

Conclusion:

As a basis of many examples of insect-plant interactions, semiochemicals are incredibly
important signals to research and utilize. As studies have shown, the complex effects of
semiochemicals are invaluable to ensure a plant’s survival against insect herbivory and to bolster
other relationships with them. However, they are also able to become a boon not only to plants,
but also to humans, who are able to manufacture semiochemicals in order to manipulate insects
much more safely than conventional methods. Due to their significant utility as tools,
semiochemicals can no longer be ignored as solutions to the multitude of problematic insect-
plant situations

References

Agelopoulos N, Birkett MA, Hick AJ, Hooper AM, Pickett JA, Pow EM, Smart LE,
Smiley DWM, Wadhams LJ, Woodcook CM (1999) Exploiting semiochemicals in insect
control. Pesticide Science 55:225-235.

Bernasconi ML, Turlings TCJ, Ambrosetti L, Bassetti P, Dorn S (1998) Herbivore-
induced emissions of maize volatiles repel the corn leaf aphid, Rhopalosiphum maidis.
Entomologia Experimentalis et Applicata 87:133-142.

Beyaert I, Hilker M (2014) Plant odour plumes as mediators of plant-insect interactions.
Biol. Rev. 89:68-81.

Cook SM, Khan ZR, Pickett JA (2006) The use of push-pull strategies in integrated pest
management. Annu. Rev. Entomol. 52:375-400.

Dettner K, Liepert C (1994) Chemical mimicry and camouflage. Annu. Rev. Entomol.
39:129-154.

Landolt PJ, Phillips TW (1997) Host plant influences on sex pheromone behavior of
phytophagous insects. Annu. Rev. Entomol. 42:371-391.

Moraes CM, Mescher MC, Tumlinson JH (2001) Caterpillar-induced nocturnal plant
volatiles repel conspecific females. Nature 410:577-580.

Your Name

Reddy GVP, Guerrero A (2004) Interactions of insect pheromones and plant
semiochemicals. Trends in Plant Science 9:253-261.

Stowe MK, Turlings TCJ, Loughrin JH, Lewis WJ, Tumlinson JH (1995) The chemistry
of eavesdropping, alarm, and deceit. Proc. Natl. Acad. Sci. USA 92:23-28.

Witzgall P, Kirsch P, Cork A (2010) Sex pheromones and their impact on pest
management. J Chem Ecol 36:80-100.

Steps in the Preparation of an Annotated Bibliography

Outline:

This outline is an overview of the paper that outlines the scope of the paper, describing the topics

to be covered and the order. This outline will contain all the detail you need to write a complete

paper. If you prepare your outline correctly, it should be almost as long as the actual paper

(don’t freak out, that’s a good thing). Below is a general example of what your outline should

look like, except that yours will be much longer.

Make sure that your outline includes the following:

1. Topic sentences for each section (marked by Roman numerals)

2. Subtopics:

• Major key points

3. References for each point that you intend to cover.

Example

Title

• Should inform the reader of what the paper is about.

• When constructing a title, choose informative over cute.

I. Introduction: Topic sentence that states the basic idea/premise of your paper. For
example, if you are examining the role of neuropeptides in parental behavior, your first

sentence might introduce parental behavior and it’s significance in species survival

(Reference- see format for in-text citations).

a. Introduce your system

i. Define the system, its function, types present, etc. (Reference- see format

for in-text citations)

ii. If you are comparing organisms, briefly introduce the organisms

(Reference- see format for in-text citations)

b. The main focus of the paper

c. Provide the scientific reasoning as to why part b is interesting

d. Be brief and concise

II. Main Body. Topic sentence focusing on your major points (Reference- see format for in-
text citations).

a. You may choose to devote one section to describe the behavior/ ecology /

scientific relevance or problem that you are focusing on.

i. Background information

1. Generalities of the taxon the species belongs to

2. Behavior and ecology of the species

3. Scientific relevance of the species (e.g. research breakthroughs that

have been possible thanks to this species)

4. Etc. (Reference- see format for in-text citations)

5. Include as many sections as you deem necessary to cover your

main ideas.

III. Conclusion
a. Summary

b. Significance

IV. References (see reference instructions)

JosephMartinez

10/16/2020

Topic in Ecology

Title: Review of The Impact of Climate Change an Wildfires and It’s Ecological Ramifications

I. Introduction: This section will focus on introducing and providing background for wildfires

(and its significance ecologically). The introduction will also introduce the concept of

climate changes as an amplifying force for intense wildfires in order to set up the

structure for the rest of the review paper.

A. Wildfires are naturally occurring phenomena that may temporarily change an

ecosystem’s composition, however most modern wildfires have had devastating

effects on ecosystems (Akaike et al., 1974).

B. Anthropogenic climate change makes intense wildfire more common (Abatzoglou

& Williams,

2016).

C. The main contributing factors are high temperatures, more severe droughts,

stronger winds, and more frequent lighting strikes- all side effects of climate

change.

II. Main Body: This section seeks to explore how wildfires naturally start and what makes

an intense wildfire, using evidence from the American West as well as the Australian

outback for a more global perspective. The main body will also break down how each

factor that contributes to intense wildfires is being amplified by climate change.

A. The Conditions Necessary for an “Intense” Wildfire- ​The main point of this section

is that wildfires needs certain conditions to thrive:

1. Wildfires need hot weather in order to take hold (Nature, 2019).

2. Wildfires need dry vegetation for “fuel” (Nature, 2019).

3. Wildfires need strong winds for oxygenation and to spread over long

distances (Nature, 2019).

4. Wildfires need a “spark” (lightning, campfire, arson, cigarette) in order to

ignite the initial flame (Nature, 2019).

B. How Climate Change is Amplifying these Conditions-​ As a follow up the the

previous section, it will be explained here how each of these conditions have

been amplified due to climate change:

1. Climate change contributes to ever hotter air and surface temperatures,

leading to “hot weather” (Hansen et al., 2006).

2. Climate change contributes to prolonged and intense droughts leading to

vast quantities of dry vegetation (Littell, Peterson, Riley, Liu, & Luce,

2016).

3. Climate change has been linked to contributing to stronger and faster

winds, which are an essential source of oxygenation and spreading

mechanisms for wildfires (Zeng et al., 2019).

4. Climate change has been linked to an increase in lightning frequency, one

of the most common “sparks” that ignite wildfires (Romps, Seeley,

Vollaro, & Molinari, 2014).

C. The Ecological Impacts Of Intense Wildfires Globally-​ In this section, the

ecological effects of wildfires will be explored in order to understand how

damaging more frequent wildfires will be in the future as climate change

progresses.

1. Wildfires release large amounts of previously trapped carbon into the

atmosphere, creating a feedback loop where wildfire emissions worsen

climate change which make wildfires more common and devastating

which leads to more emissions (van der Werf et al., 2017).

2. Wildfires produce a hydrophobic soil after its been charred and cleared of

vegetation, leading to more pronounced runoffs after rain. The result is a

large amount of contaminated water flowing into nearby bodies of water

dramatically changing their concentrations of nitrogen and phosphorus

and introducing heavy metals (Hallema et al., 2018).

3. Intense wildfires can change and reduce the regional biodiversity through

making certain areas unsuitable for some plants and animals after the

dramatic change to the environment (Stevens-Rumann et al., 2017).

III. Conclusion: In this section, the paper will attempt to tie in all the evidence and reiterate

its results in a concise way. In addition to that, it will point out the relevance of the paper

holds given current events. This section will also explain the significance and importance

of intense wildfires being caused by climate change and what we can expect in the

future as climate change progresses and wildfires become more frequent and

devastating.

IV. Citations

A. Williams, A., Abatzoglou, J., Gershunov, A., Guzman-Morales, J., Bishop, D.,

Balch, J., & Lettenmaier, D. (2019, August 04). ​Observed Impacts of

Anthropogenic Climate Change on Wildfire in California​. Retrieved October 15,

2020, from

https://agupubs.onlinelibrary.wiley.com/doi/full/10.1029/2019EF001210

B. Abatzoglou, J., & Williams, A. (2016, October 18).​ Impact of anthropogenic

climate change on wildfire across western US forests​. Retrieved October 15,

2020, from ​https://www.pnas.org/content/113/42/11770

C. LeRoy, W., Anthony LeRoy Westerling Anthony LeRoy Westerling

http://orcid.org/0000-0003-4573-0595\, Westerling, A., Anthony LeRoy Westerling

https://agupubs.onlinelibrary.wiley.com/doi/full/10.1029/2019EF001210

https://www.pnas.org/content/113/42/11770

http://orcid.org/0000-0003-4573-0595 Google Scholar Find this author on

PubMed Search for more papers by this author, One contribution of 24 to a

discussion meeting issue ‘The interaction of fire and mankind’., & Al., E.

(2016, June 05).​ Increasing western US forest wildfire activity: Sensitivity to

changes in the timing of spring​.

Retrieved October 15, 2020, from

https://royalsocietypublishing.org/doi/10.1098/rstb.2015.0178

D. Romps, D., Seeley, J., Vollaro, D., & Molinari, J. (2014, November 14).

Projected increase in lightning strikes in the United States due to global warming​.

Retrieved October 15, 2020, from

https://science.sciencemag.org/content/346/6211/851.abstract

E. Liu, Y., Stanturf, J., & Goodrick, S. (2009, October 03). ​Trends in global

wildfire potential in a changing climate​. Retrieved October 15, 2020, from

https://www.sciencedirect.com/science/article/abs/pii/S0378112709006148

F. Stevens-Rumann, C., Kemp, K., Higuera, P., Harvey, B., Rother, M., Donato, D.,

Veblen, T. (2017, December 12).​ Evidence for declining forest resilience to

wildfires under climate change​. Retrieved October 15, 2020, from

https://onlinelibrary.wiley.com/doi/abs/10.1111/ele.12889

G. Higuera, P., Abatzoglou, J., Littell, J., & Morgan, P. (n.d.). ​The Changing

Strength and Nature of Fire-Climate Relationships in the Northern Rocky

Mountains​, U.S.A., 1902-2008. Retrieved October 15, 2020, from

https://journals.plos.org/plosone/article?id=10.1371%2Fjournal.pone.0127563

H. Running, S. (2006, August 18).​ Is Global Warming Causing More, Larger

Wildfires?​ Retrieved October 15, 2020, from

https://science.sciencemag.org/content/313/5789/927

I. Westerling, A., Hidalgo, H., Cayan, D., & Swetnam, T. (2006, August 18).

Warming and Earlier Spring Increase Western U.S. Forest Wildfire Activity​.

https://royalsocietypublishing.org/doi/10.1098/rstb.2015.0178

https://science.sciencemag.org/content/346/6211/851.abstract

https://www.sciencedirect.com/science/article/abs/pii/S0378112709006148

https://onlinelibrary.wiley.com/doi/abs/10.1111/ele.12889

https://journals.plos.org/plosone/article?id=10.1371%2Fjournal.pone.0127563

https://science.sciencemag.org/content/313/5789/927

Retrieved October 15, 2020, from

https://science.sciencemag.org/content/313/5789/940

J. Dennison, P., Brewer, S., Arnold, J., & Moritz, M. (2014, April 25).​ Large

wildfire trends in the western United States, 1984–2011​. Retrieved October 15,

2020, from ​https://agupubs.onlinelibrary.wiley.com/doi/10.1002/2014GL059576

K. Littell, J., McKenzie, D., Peterson, D., & Westerling, A. (2009, June 01).

Climate and wildfire area burned in western U.S. ecoprovinces, 1916–2003​.

Retrieved October 15, 2020, from

https://esajournals.onlinelibrary.wiley.com/doi/10.1890/07-1183.1

L. Littell, J., Peterson, D., Riley, K., Liu, Y., & Luce, C. (2016, April 19).​ A review of

the relationships between drought and forest fire in the United States​. Retrieved

October 15, 2020, from ​https://onlinelibrary.wiley.com/doi/abs/10.1111/gcb.13275

M. Harvey, B. (2016, October 15).​ Human-caused climate change is now a key

driver of forest fire activity in the western United States​. Retrieved October 15,

2020, from ​https://www.pnas.org/content/113/42/11649

N. Akaike, H., CD. Allen, D., BJ. Bentz, J., C. Boisvenue, S., Breiman, L., DD.

Breshears, N., RH. Waring, S. (1974, January 01).​ Forest ecosystems,

disturbance, and climatic change in Washington State, USA​. Retrieved October

15, 2020, from ​https://link.springer.com/article/10.1007/s10584-010-9858-x

O. Zeng, Z., Ziegler, A., Searchinger, T., Yang, L., Chen, A., Ju, K., . . . Wood, E.

(2019, November 18).​ A reversal in global terrestrial stilling and its implications

for wind energy production​. Retrieved October 15, 2020, from

https://www.nature.com/articles/s41558-019-0622-6

P. Hallema, D., Sun, G., Caldwell, P., Norman, S., Cohen, E., Liu, Y., . . . McNulty,

S. (2018, April 10).​ Burned forests impact water supplies​. Retrieved October 15,

2020, from ​https://www.nature.com/articles/s41467-018-03735-6

https://science.sciencemag.org/content/313/5789/940

https://agupubs.onlinelibrary.wiley.com/doi/10.1002/2014GL059576

https://esajournals.onlinelibrary.wiley.com/doi/10.1890/07-1183.1

https://onlinelibrary.wiley.com/doi/abs/10.1111/gcb.13275

https://www.pnas.org/content/113/42/11649

https://link.springer.com/article/10.1007/s10584-010-9858-x

https://www.nature.com/articles/s41558-019-0622-6

https://www.nature.com/articles/s41467-018-03735-6

Q. Van der Werf, G., Randerson, J., Giglio, L., Van Leeuwen, T., Chen, Y., Rogers,

B., . . . Kasibhatla, P. (2017, September 12). ​Global fire emissions estimates

during 1997–2016​. Retrieved October 15, 2020, from

https://essd.copernicus.org/articles/9/697/2017/

R. Hansen, J., Sato, M., Ruedy, R., Lo, K., Lea, D., & Medina-Elizade, M.

(2006, September 26).​ Global temperature change​. Retrieved October 15, 2020,

from ​https://www.pnas.org/content/103/39/14288

S. The complexities of wildfires​. (2019, January 30). Retrieved October 15, 2020,

from https://www.nature.com/articles/s41561-019-0311-0

https://essd.copernicus.org/articles/9/697/2017/

https://www.pnas.org/content/103/39/14288

Invasive reptile species of Florida.

I. Introduction:

The problem of invasive species is relevant all over the world. In the state of Florida, this

issue is especially acute due to the hospitable climate that due to warm temperatures and

increases air moisture makes it incredibly easy for newly introduced species to thrive. The

invasive species may potentially damage the environment in many ways, human economy,

health, safety, and negatively impact native species. Among other animals, the reptiles comprise

quite a long list of the invasive species in Florida: Argentine black and white tegu (Tupinambis

merianae), black spiny-tailed iguana (Ctenosaura similis), brown anole (Anolis sagrei), the

Burmese python (Python bivittatus), common house gecko (Hemidactylus frenatus), green

iguana (Iguana iguana), Mediterranean gecko (Hemidactylus turcicus), Nile Monitor (Varanus

niloticus). In this review paper, some of the most invasive reptile species in the state of Florida

will be discussed. Their history, origin, and impact on native species and the environment will be

addressed. In final part will focus on the investigation of steps and measures taken or planned to

be taken in the future to reduce or eliminate these species from the state.

II. History and origin of reptile invasive species in Florida.

In this section, the history of invasive reptiles will be discussed, the ways of the invasion

as well as some aspects of the dispersal. There are multiple ways the species may be introduced.

It may be a natural or human-facilitated event, accidental or deliberate, such as pet trade or zoo

escape. The key aspects of their success will be presented (Engeman et al., 2011).

i. Green iguana (Krysko et al., 2007)

ii. Nile monitor (Enge et al., 2004), (Wood et al., 2016).

iii. Burmese python (Wilson et al., 2011).

iv. Black and White Tegu (Pernas et al., 2012).

v. Human mediated dispersal on the example of common house gecko (Short &

Petren, 2011), (Muller et al., 2020).

III. Impact on native species.

This section will cover particular examples of some invasive reptiles on native species. The

introduction of the genetic variation as one of the effects will be introduced. Besides, the

interesting effect of the attempt of their extraction will be considered.

i. Brown anole effect on native lizard species in Florida (Campbell, 2000).

ii. Birds predation by Burmese python (Dove et al., 2011).

iii. The decline of the tree snail due to predation by green iguana in Key Biscayne, Florida

(Townsend et al., 2005).

iv. Discussion of the genetic paradox of the increase in genetic variation during the invasion

of brown anole (Kolbe et al., 2004).

v. Removal of some invasive species such as green iguanas may negatively affect their

predators, such as gray fox and raccoons (Meshaka et al., 2007).

IV. Influence on the environment.

i. It is important to understand how certain species such as the Burmese python, select the

locations of their habitat. It would allow identifying the consequences, areas most

affected, and the scale and direction of the possible spread. (Walters et al., 2016)

ii. Burmese python as hosts for native mosquito communities may introduce negative

implications such as an increase in mosquito population and transmission of mosquito-

vectored pathogens (Reeves et al., 2018).

V. Economic effects.

This section will touch upon some economic implications caused by the invasion of reptiles

using the green iguana as an example.

i. Associated costs and policy implications due to green iguana invasion and consequent

infrastructural damages may be counted in billions of dollars. (Sementelli et al., 2008).

VI. Invasive reptile species management.

Steps currently that are being taken to reduce the populations of invasive reptile species will

be discussed in this section as well as some proposed solutions.

ii. Annual event National Invasive Species Awareness Week that takes place since 2010

raises the awareness and seeks for solutions (NISAW).

iii. Overview of Burmese python management steps taken in 2013 (Mazzotti, 2016).

VII. Conclusion

The invasion of the species into the non-native environment may not always have

negative consequences; in some rare cases, it might be even beneficial. However, in the case of

Florida, the invasive reptile species bring a lot of harm to not only the local environment and

ecology but also are quite costly to be managed to reduce the destructive effects of their

presence. It is a task of high importance to preserve the abundant natural habitat of the southern

state and limit the new invasions on the legislative level. This paper reviewed some of the most

abundant and damaging species, the ways and history of their introduction, the effects they

caused, and the possible ways of the control for their population numbers and spreading.

References

Campbell, T. S. (2000). Analysis of the effects of an exotic lizard (Anolis sagrei) on a

native lizard (Anolis carolinensis) in Florida, using islands as experimental units. PhD thesis,

Univ. Tennessee

Dove, C. J., Snow, R. W., Rochford, M. R., Mazzotti, F. J. (2011). Birds Consumed by

the Invasive Burmese Python (Python Molurus Bivittatus) in Everglades National Park, Florida,

USA. The Wilson Journal of Ornithology, 123(1), 126.

Enge, K. M., Krysko, K. L., Hankins, K. R., Campbell, T. S., & King, F. W. (2004).

Status of the Nile monitor (Varanus niloticus) in southwestern Florida. Southeastern Naturalist,

3(4), 571-582.

Wood, J. P., Dowell, S. A., Campbell, T. S., Page, R. B. (2016) Insights into the

Introduction History and Population Genetic Dynamics of the Nile Monitor (Varanus niloticus)

in Florida, Journal of Heredity, Volume 107, Issue 4, July 2016, Pages 349–362, https://doi-

org.ezproxy.fiu.edu/10.1093/jhered/esw014

https://doi-org.ezproxy.fiu.edu/10.1093/jhered/esw014

https://doi-org.ezproxy.fiu.edu/10.1093/jhered/esw014

Townsend, J. H., Slapcinsky, J., Krysko, K. L., Donlan, E. M., Golden, E. A. (2005).

Predation of a Tree Snail Drymaeus multilineatus (Gastropoda: Bulimulidae) by Iguana iguana

(Reptilia: Iguanidae) on Key Biscayne, Florida. Southeastern Naturalist, 4(2), 361.

Kolbe, J., Glor, R., Rodríguez Schettino, L. (2004). Genetic variation increases during

biological invasion by a Cuban lizard. Nature 431, 177–181 https://doi.org/10.1038/nature02807

Krysko, K. L., Enge, K. M., Donlan, E. M., Seitz, J. C. (2007) Distribution, Natural

History, and Impacts of the Introduced Green Iguana in Florida, Iguana: Conservation, Natural

History, and Husbandry of Reptiles, International Reptile Conservation Foundation, 14 (3): 142–

151

Mazzotti F. J., Rochford M., Vinci J., Jeffery B. M., Eckles J. K., Dove C., Sommers K.

P. (2016). Implications of the 2013 Python Challenge® for Ecology and Management of Python

Molorus bivittatus (Burmese Python) in Florida. Southeastern Naturalist, 15, 63.

Meshaka, W., Bartlett, R.,Smith, H. (2004) Colonization success by Green Iguanas in

Florida. Iguana: Journal of the International Iguana Society. State Museum of Pennsylvania,

Zoology and Botany. 11 (3): 154–161.

Muller, B. J., Andrews, R. M., Schwarzkopf, L., Pike, D. A. (2020). Social context alters

retreat- and nest-site selection in a globally invasive gecko, Hemidactylus frenatus. Biological

Journal of the Linnean Society, 129(2), 388–397. https://doi-

org.ezproxy.fiu.edu/10.1093/biolinnean/blz188

Reeves, L. E., Krysko, K. L., Avery, M. L., Gillett-Kaufman, J. L., Kawahara, A. Y.,

Connelly, C. R., & Kaufman, P. E. (2018). Interactions between the invasive Burmese python,

Python bivittatus, and the local mosquito community in Florida, USA. PloS one, 13(1),

e0190633. https://doi.org/10.1371/journal.pone.0190633

Sementelli, A., Smith, H. T., Meshaka, W., Engeman, R. M. (2008). Just Green Iguanas?

The Associated Costs and Policy Implications of Exotic Invasive Wildlife in South Florida.

Public Works Management Policy 2008; 12; 599.DOI: 10.1177/1087724X08316157

https://doi.org/10.1038/nature02807

https://doi-org.ezproxy.fiu.edu/10.1093/biolinnean/blz188

https://doi-org.ezproxy.fiu.edu/10.1093/biolinnean/blz188

https://doi.org/10.1371/journal.pone.0190633

Short, K. H., Petren, K. (2011). Multimodal dispersal during the range expansion of the

tropical house gecko Hemidactylus mabouia. Ecology and evolution, 1(2), 181–190.

https://doi.org/10.1002/ece3.18

Walters, T. M., Mazzotti, F. J., Fitz, C. (2016). Habitat Selection by the Invasive Species

Burmese Python in Southern Florida. Journal of Herpetology, 50(1), 50.

Pernas, T., Giardina, D. J., McKinley, A., Parns, A., Mazzotti, F. J. (2012). First

Observations of Nesting by the Argentine Black and White Tegu, Tupinambis merianae, in

South Florida. Southeastern Naturalist, 11(4), 765.

Willson, J. D., Dorcas, M. E., Snow, R. W. (2011). Identifying plausible scenarios for the

establishment of invasive Burmese pythons (Python molurus) in Southern Florida. Biological

Invasions, 7, 1493.

https://doi.org/10.1002/ece3.18