“Phosphorus availability and microbial
driven decomposition in oligotrophic
seagrass sediments”
Matthew William Fraser
B.Sc. (Hons.) Marine Science (UWA)
This thesis is presented for the degree of Doctor of Philosophy at the
School of Plant Biology, University of Western Australia
2016
Phosphorus cycling in seagrass sediments
Abstract
Phosphorus (P) availability has long been considered to be limiting in many seagrass
communities. However, seagrass meadows are often highly productive in oligotrophic
waters despite extremely low nutrient concentrations. Seagrass sediments, in contrast, are
often enriched in organic matter (OM) that could potentially be a major source of
nutrients, particularly of P, following microbial degradation. The main aim of this thesis
was to investigate how seagrasses productivity was supported in Shark Bay – an
embayment with extremely low soluble reactive phosphorus (SRP) in the water column
but with abundant seagrass meadows. I address this key question by focusing on the
importance of OM decomposition for P cycling and plant uptake in oligotrophic seagrass
sediments.
The research focused on Shark Bay, a World Heritage Site in Western Australia
dominated by seagrasses that form the foundation of this important ecosystem.
I
first document a major pulse of OM inputs to Shark Bay sediments following widespread
defoliation and dieback of the dominant seagrass Amphibolis antarctica in Shark Bay
driven by a marine heatwave and flood event. Given the frequency and magnitude of
extreme events are predicted to increase, such events will likely influence biogeochemical
cycling in Shark Bay and similar seagrass ecosystems in the near future. I then
investigated the effects of seagrass-derived OM on sediment biogeochemistry and
microbial communities, and seagrass seedling health in controlled tank systems.
Seagrass-derived OM changed the sediment microbial community and significantly
increased hydrolytic enzyme expression, particularly phosphatase and glucosidase
enzymes. OM enrichment also correlated with decreased N and P content of living
seagrass leaves, likely caused by increased microbial competition for nutrients decreasing
plant uptake or higher aboveground biomass diluting nutrient impacts. Seagrass derived
OM has previously been considered a refractory pool of nutrients, yet here I show its
deposition into sediments significantly alters belowground conditions.
I next focused on the specific effects of OM enrichment on P cycling across Shark Bay
sediments. I characterised the different fractions of sedimentary phosphorus across the
salinity gradient in Shark Bay sediments, including examining the organic P pool to the
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Fraser (2016)
compound level using 31P-NMR (nuclear magnetic resonance). I show that organic P
represents a substantial (16-40%) proportion of total P in Shark Bay sediments, and that
this pool is strongly correlated with OM content in sediments. The only identifiable
organic P compounds detected using 31P-NMR were phosphate monoesters. The percent
organic P was highest at sites with low water SRP, representing a potential source of P at
these sites.
I then examined how sediment microbial community composition and functioning relates
to patterns in biogeochemical processes and seagrass nutrient status in Shark Bay. I thus
characterised the taxonomic and functional diversity of microbial communities within
seagrass sediments along a gradient of decreasing water SRP using metagenomic
sequencing to better understand P dynamics in the sediments. Genes encoding for
phosphorus metabolism increased at low water SRP sites and also correlated to an
increase in phosphatase activity. The increase in both the presence of P metabolism gene
abundance and the activity of phosphatases suggests that microbial recycling helps
support seagrass P demand in Shark Bay through mineralization or phosphate monoesters.
Notably, seagrass microbial communities were high in gene abundance for sulfur and
phosphorus metabolism when compared with other aquatic and terrestrial ecosystems.
I also sought to investigate the extent to which microbial turnover of organic P plays in
seagrass uptake of P across the salinity and P availability gradient, using stable oxygen
isotope composition of phosphates (δ18OPO4) as a tracer for biological cycling of P.
However, my attempts to measure δ18OPO4 in seagrass and sediment samples were
unsuccessful, despite successfully processing and obtaining correct δ18OPO4 for
international isotope standards. Extremely low phosphate concentrations and high salinity
(>50‰) prevented successful extraction of pure silver phosphates and led to incorrect
δ18OPO4 for Shark Bay seagrass and sediment samples.
Overall, this thesis shows that seagrass derived OM in sediments cannot be viewed as a
passive pool of nutrients in seagrass sediments, and the turnover of OM in seagrass
meadows is a potential source of P in oligotrophic seagrass ecosystems. The turnover of
OM in strongly oligotrophic seagrass sites is dependent on highly specialised and adapted
microbial communities that have a key role in driving mineralization processes, showing
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Phosphorus cycling in seagrass sediments
that interactions between seagrasses and sediment microbial communities are a critical
knowledge gap within seagrass ecology, and must be further characterised given their
impacts on nutrient cycling and primary productivity.
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Fraser (2016)
Table of Contents
Abstract ............................................................................................................................. i
Table of Contents ........................................................................................................... iv
Acknowledgements........................................................................................................ vii
Statement of candidate contribution and list of publications ..................................... x
Thesis Structure ............................................................................................................ xii
Chapter 1: General introduction ................................................................................... 1
The productivity of seagrass ecosystems ...................................................................... 2
Marine sediments and the phosphorus cycle................................................................. 3
Organic matter and nutrient cycling in near coastal ecosystems .................................. 5
Phosphorus availability and organic matter turnover ................................................... 5
Microbial ecology – engines that drive biogeochemical cycles ................................... 7
Shark Bay: a model oligotrophic ecosystem ................................................................. 9
Thesis aims and structure ............................................................................................ 10
References ................................................................................................................... 12
Chapter 2: Extreme climate events lower resilience of foundation seagrass at edge
of biogeographical range .............................................................................................. 20
Summary ........................................................................................................................ 21
Introduction ................................................................................................................. 22
Methods ....................................................................................................................... 25
Study site ............................................................................................................................. 25
Spatial extent of Amphibolis defoliation ............................................................................. 25
Data analysis ...................................................................................................................... 26
Results ......................................................................................................................... 28
Extreme temperature and rainfall ....................................................................................... 28
Response of seagrasses to flooding and temperature events .............................................. 29
Recovery of seagrass meadows ........................................................................................... 31
Discussion ................................................................................................................... 32
Acknowledgments ....................................................................................................... 38
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References ................................................................................................................... 38
Chapter 3: Seagrass derived organic matter influences biogeochemistry, microbial
communities, and seedling biomass partitioning in seagrass sediments .................. 43
Abstract ....................................................................................................................... 44
Introduction ................................................................................................................. 45
Methods ....................................................................................................................... 47
Seedling collection ............................................................................................................. 47
Experimental design ........................................................................................................... 47
Post-harvest measurements ................................................................................................ 48
Sediment biogeochemistry .................................................................................................. 49
Microbial molecular community identification .................................................................. 49
Statistical analysis .............................................................................................................. 50
Results ......................................................................................................................... 51
Sediment biogeochemistry .................................................................................................. 51
Microbial community structure .......................................................................................... 53
Seagrass biomass ............................................................................................................... 54
Seagrass nutrient concentrations ....................................................................................... 56
Discussion ................................................................................................................... 57
Acknowledgements ..................................................................................................... 62
References ................................................................................................................... 62
Chapter 4: Organic phosphorus accumulates in calcareous seagrass sediments with
increasing salinity .......................................................................................................... 68
Abstract ....................................................................................................................... 69
Introduction ................................................................................................................. 70
Methods ....................................................................................................................... 72
Site description & experimental design.............................................................................. 72
Sediment cores ................................................................................................................... 73
Sediment chemical and physical characterization ............................................................. 74
Sequential P extraction of sediments.................................................................................. 74
Solution 31P-NMR spectroscopy ......................................................................................... 75
Data analysis ...................................................................................................................... 75
Results ......................................................................................................................... 76
General sediment variation along salinity gradient........................................................... 76
Sediment P fractions........................................................................................................... 76
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Solution 31P-NMR spectroscopy.......................................................................................... 82
Discussion ................................................................................................................... 83
Acknowledgments ....................................................................................................... 88
References ................................................................................................................... 88
Chapter 5: Taxonomic and functional structure of microbial communities in
seagrass sediments across salinity and phosphorus gradients .................................. 93
Methods ..................................................................................................................... 106
Sampling sites and experimental design ........................................................................... 106
Enzymes ............................................................................................................................ 106
Microbial community identification.................................................................................. 107
Metagenomic annotation and statistical analysis ............................................................. 107
References ................................................................................................................. 109
Chapter 6: Tracking phosphorus cycling in seagrass meadows using oxygen
isotopes: Methodological challenges and future approaches .................................. 112
Introduction ............................................................................................................... 113
Methods ..................................................................................................................... 114
Sampling sites and experimental design ........................................................................... 114
Results ....................................................................................................................... 117
Discussion ................................................................................................................. 117
References ................................................................................................................. 122
Chapter 7: General discussion: organic matter cycling and the phosphorus
biogeochemistry of Shark Bay ................................................................................... 125
Summary of main findings of this research .............................................................. 127
Sediment organic matter in seagrass meadows ......................................................... 127
Seagrass ecosystems – a new focus on organic P ..................................................... 129
Extreme events: an emerging threat to seagrass meadows ....................................... 130
Microbial ecology of seagrass sediments.................................................................. 132
A new assessment of the biogeochemistry of Shark Bay ......................................... 134
Conclusion ................................................................................................................ 136
References ................................................................................................................. 136
Appendices ................................................................................................................... 143
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Acknowledgements
I must begin by thanking my ever patient supervisors – Gary Kendrick, Pauline Grierson,
and Grzegorz Skrzypek. Gary, you have always been extremely generous to me, have
looked out for my best interests and allowed me the freedom to explore my own interests
without wandering too far. You have given me opportunities beyond my PhD work that
most postgraduate students would only dream off, and I am deeply grateful for your belief
in my ability. Pauline, as a fellow phosphorus biogeochemistry and coffee tragic you have
provided me with wisdom as I often questioned my own research direction. No matter
how much I felt I knew coming into a meeting with you, you always managed to give me
a fresh perspective and run off to read more papers or run more analyses. Greg, even
though our foray into δ18O-PO4 analysis didn’t quite work out as we had hoped, I have
learned much from you during my PhD, and you have always been more than generous
with your time. Above all, I will be forever grateful with the respect that all my
supervisors treated me with – not merely a student-supervisor relationship, but as genuine
collaborations built upon trust and respect. I aspire to take the lessons you have taught
with me should I ever get the opportunity to be the supervisor instead of the student.
There have been many fantastic collaborators that I have worked with on various aspects
of this thesis, which have undoubtedly improved its breadth and content. Their knowledge
and expertise inspired me to strive to improve my own research. Bonnie and Deirdre –
thanks for introducing me to the small but mighty world of microbes! I loved delving into
this work with you, and I’m sure that there will be many further microbial projects in the
future. Di – without the countless hours you have put into seagrass research at UWA then
I would not have had this opportunity. I am deeply grateful for this, and for you
introducing me to the wonderful place that is Shark Bay. And a special thanks to Renae
and John – you are both lifelong friends – working together is just a bonus! I likely would
be doing something very different if not for you two, you prevented me from heading
over to the dark side and sticking with benthic research. There have also been many
technical staff that have assisted me on my journey, with a special thanks to Andrea
Zavala-Perez, Kate Bowler, Doug Ford and Lindsey Byrne.
I was supported by an Australian Postgraduate Award with a UWA Safety-Net Top-Up
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Fraser (2016)
from the University of Western Australia. Funding for the project was primarily provided
by a NHT-II Caring for Country Grant. Additional funding was provided by the School
of Plant Biology (UWA), the Ray Hart Memorial Scholarship, and the Australian
Geographic Society. I thank the Department of Parks and Wildlife for providing permits
that allowed me to complete the research presented here. I’d also like to acknowledge the
School of Plant Biology – the research culture in the School is fantastic and a credit to all
the staff involved that make the PhD experience so rewarding for postgrads like me.
I’ve been lucky enough to be part of two fantastic research groups during my PhD at
UWA. The Seagrass Research Group is full of like-minded people that love nothing more
than being on the water (and finishing fieldwork days watching the sun go down with a
brew!). I couldn’t have completed the fieldwork without you guys, and the many meetings
and writing sessions we’ve had have helped shape the way I think about research. Uncle
John, Auntie Nae, Andrea, Leo, Liz, Anne, Marion, Ylva, Luke, Napo, Belinda, Dan,
Sam, Sahira (thanks for making my first supervisory role so easy!) – you’ve all been great.
And thanks for putting up with all of us Gary! To ERGo: you became my home away
from home while I was nearing the end of the thesis, which just so happened to be the
perfect time to find a group that was always interested in stopping for a coffee break.
Pauline, Greg, Kate, Sara, Doug, Alex, Ali, Andre, Rachel, Tegan, Jordan, and Caroline
– thank-you all for putting up with me (and all the baby talk) while I was nearing the end
of my thesis.
Finally, I wish to thank my family. You are the most important thing in the world to me,
and your love and care remind me not to get bogged down in work. Mum and Dad, thank
you for providing a loving and stable environment for me to grow up in, sparking off my
interest and science and catching me whenever I fell. Scott and CJ – I couldn’t ask for
better siblings. Scott and Annie, thanks for convincing us all to come to Australia in the
first place, it is without a doubt the best move we all made. And Tara – no words can
express how grateful I am for your love. No matter what bumps I encountered along the
way, you were always there to remind me to believe in myself and support me through
thick and thin. You are the love of my life and I wouldn’t choose to be with any other
person on our new adventure. And to my two newborn sons, Harper and Cooper, you
provided me with the motivation to finish this thesis. Having identical twins within a
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Phosphorus cycling in seagrass sediments
month of your PhD submission may not seem like a good idea, but I wouldn’t have
changed it for the world! Like your mum, you both mean the world to me. This thesis is
dedicated to my family, I love you all
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Fraser (2016)
Statement of candidate contribution and list of publications
The following chapters have been published as scientific manuscripts:
Chapter 2
Fraser MW, Kendrick GA, Statton J, Hovey RK, Zavala Perez A, Walker DI (2014)
Extreme climate events lower resilience of foundation seagrass at edge of
biogeographical range. Journal of Ecology 102:1528–1536. doi: 10.1111/13652745.12300
Chapter 3
Fraser MW, Statton J, Hovey RK, Laverock B, & Kendrick GA (2015). Seagrass derived
organic matter influences biogeochemistry, microbial communities, and seedling biomass
partitioning in seagrass sediments Plant and Soil (2015). doi:10.1007/s11104-015-27210
In addition, two other chapters have been prepared for submission to scientific journals:
Chapter 4
Fraser MW, Kendrick GA, Skrzypek G, Byrne L, Grierson PF. Organic phosphorus
accumulates in calcareous seagrass sediments with increasing salinity Prepared for:
Limnology & Oceanography. Submitted.
Chapter 5
Fraser MW, Gleeson DB, Grierson PF, Laverock B, Kendrick GA Taxonomic and
functional structure of microbial communities in seagrass sediments across salinity and
phosphorus gradients Prepared for: Nature Microbiology. In preparation
I conceived all research questions (in consultation with supervisors), designed the
research program, was responsible for the lab work, data analysis and interpretation, as
well as the preparation of each manuscript. The thesis is my own work, except where
explicitly stated otherwise. The contribution of co-authors to each chapter is described
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in the preamble to each chapter. All co-authors consent to this work being included in this
thesis. Co-authors primarily provided advice on initial research direction, specific advice
on experimental methods and provided feedback on manuscript drafts.
Signed:
Matthew Fraser (Candidate)
Gary Kendrick (Coordinating Supervisor)
September 2016
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Fraser (2016)
Thesis Structure
This thesis has been prepared as a ‘series of papers’ as per Regulation 1.3.1.33 (1) of the
University of Western Australia’s General Provisions for Research Higher Degrees (by
Thesis). Five experimental chapters (two published, two to be submitted, one a method
commentary on oxygen isotope composition of phosphates) are presented, preceded by a
General Introduction to place the research in context. A final General Discussion
synthesises the overall findings and discusses the implications of the research. The five
experimental chapters are written as “stand-alone” manuscripts. Thus, there is some
repetition (particularly in materials and methods sections), although every effort has been
made to keep this repetition to a minimum. Each experimental chapter is reproduced
verbatim, including individual Chapter Abstracts, with the following exceptions:

Author addresses have been omitted

Figure and table numbering has been slightly altered to include chapter numbers

Minor changes have been made to the formatting of each chapter to ensure
consistency across the thesis.
References are included at the end of each chapter, in the style of the journal where the
research is published/to be submitted to.
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Chapter 1: General introduction
The main aim of this thesis is to investigate how seagrasses productivity is supported in
oligotrophic environments, specifically focused on the UNESCO World Heritage listed
Shark Bay region of Western Australia. In particular, I examined the importance of
decomposition and mineralization of organic matter for phosphorus (P) cycling in
oligotrophic seagrass sediments. This thesis also emphasizes the key roles that specialist
microbial communities in the sediments play in determining organic matter turnover.
How highly productive and frequently diverse ecological communities thrive in low
nutrient (oligotrophic) ecosystems has been a central ecological question since the
voyages of Darwin (Darwin 1842). In marine ecosystems, this apparent paradox has been
most often investigated for coral reefs, which boast high primary productivity and species
richness despite existing in some of the most nutrient-poor waters in the world (Smith
1984a; Hatcher 1990; Wyatt et al. 2012). Recently, new methods such as stable isotope
tracing have shed light on some of the mechanisms that contribute to higher productivity
that might be predicted on the basis of climate and nutrient availability in the water
column, including highly efficient nutrient recycling pathways existing in coral reef
ecosystems (de Goeij et al. 2013). In comparison, mechanisms driving high productivity
in other coastal marine ecosystems are understudied. Seagrass meadows are also highly
productive in oligotrophic waters, but we do not fully understand how or why (Romero
et al. 2006). Given that seagrasses can take up nutrients from sediments through their
roots, the cycling of sediment nutrient pools may also maintain seagrass productivity in
oligotrophic ecosystems. The main aim of this thesis is to investigate how seagrasses
productivity is supported in oligotrophic environments. I address this key question by
looking specifically at the importance of decomposition of organic matter for phosphorus
(P) cycling in oligotrophic seagrass sediments.
In this introductory chapter, I provide an overview of drivers of seagrass productivity in
oligotrophic environments. I then summarise the importance of phosphorus cycling in
seagrass meadows. I also provide an overview of the current understanding of organic
matter dynamics in coastal environments, before discussing if the turnover of organic
matter in seagrass sediments driven by sediment microorganisms may be important in
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supplementing P availability to seagrasses. This background thus provides context for the
experimental rationale that is developed in Chapters 2-5 of this thesis.
The productivity of seagrass ecosystems
Seagrasses are marine angiosperms that are the dominant primary producers in many
coastal ecosystems, often with high production rates comparable to coral reefs and
tropical forests (Duarte & Chiscano 1999). However, some of the most productive
seagrass meadows exist in waters with extremely low available nutrient concentrations.
For example, Posidonia oceanica (L.) growing in the in Mediterranean Sea, Thalassia
testudinum Banks ex König in the Caribbean, and Amphibolis antarctica (Labill.) Sonder
et Aschers in Shark Bay all form productive seagrass meadows in nutrient depleted waters
(Walker, Kendrick & McComb 1988; Marbà, Duarte & Cebrian 1996; Hemminga, Marbà
& Stapel 1999). Indeed, seagrass cover often declines when nutrient availability increases
due to increased competition from algae that results in light reduction (Ferdie &
Fourqurean 2004). Clearly, seagrasses can thrive in nutrient poor environments such as
Shark Bay, which suggests functional adaptation at both the plant and ecosystem level.
Seagrasses possess traits that allow them to efficiently use nutrients. For example,
seagrasses can translocate nutrients from old leaves prior to senescence (Hemminga et al.
1999). Some seagrasses also release organic anions that increase nutrient availability,
particularly of phosphates, in surrounding calcareous sediments through dissolution
(Long et al. 2008). However, such adaptations are unlikely to fully explain high seagrass
productivity in oligotrophic environments given their high nutrient requirements
(Hemminga et al. 1999). As such, mechanisms other than individual plant traits must
contribute to the high primary productivity in oligotrophic seagrass ecosystems.
Seagrasses, like terrestrial angiosperms, interact with sediments through their welldeveloped root systems. These root systems place seagrasses in a unique position for
marine primary producers in that they are able to directly access sediment nutrients
through root uptake, as well as take up nutrients from the water column through leaves
(Stapel et al. 1996; Nielsen, Koch & Madden 2007). Given seagrasses appear to take up
adequate nutrients even when water column concentrations are low suggests that
belowground processes may play a key but as yet poorly defined role in maintaining the
productivity of seagrass meadows. Furthering our understanding of below-ground
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Phosphorus cycling in seagrass sediments
processes in seagrass sediments will therefore improve our understanding of nutrient
cycling and maintenance of primary productivity in these ecosystems.
Given seagrass meadows are recognized as important marine carbon sinks it is critical to
understand the major constraints on their carbon sequestration (Kennedy et al. 2010;
Mcleod et al. 2011; Fourqurean et al. 2012a; Laliberte, Zemunik & Turner 2014),
including the limitation or otherwise of particular nutrients. Seagrasses are declining
globally, in large part attributable to direct and indirect impacts of nutrient pollution from
adjacent coastal development (Orth et al. 2006). Seagrasses growing in oligotrophic
environments are particularly susceptible to changes in labile P concentrations. For
example, rock phosphate additions of 0.12 g P m-2 d-1 to nearshore seagrass beds in
Florida Bay resulted in increases in the relative abundance of macroalgae and epiphytes,
thereby reducing seagrass growth rates (Darwin 1842; Ferdie & Fourqurean 2004).
Recovery from the effects of nutrient enrichment can be slow in oligotrophic
embayments, with seagrass community structure still altered two decades after
enrichment events in Florida Bay (Smith 1984a; Hatcher 1990; Herbert & Fourqurean
2008; Wyatt et al. 2012). Increases in P inputs are thus considered the biggest threat to
seagrass communities in sub-tropical embayments around the world, including Shark Bay
in Western Australia and Florida Bay in the southern United States (Kendrick et al. 2012;
de Goeij et al. 2013). It is thus increasingly evident that a comprehensive understanding
of the mechanisms and controls on P cycling is required if we are to fully understand
primary productivity and benthic ecology in seagrass ecosystems, and is also fundamental
to successful management of seagrass meadows and coastal zones more broadly.
Marine sediments and the phosphorus cycle
Marine sediments are a major sink and source of P in coastal ecosystems, and are
therefore an important component of the overall P cycle. Phosphorus is present in many
different forms within marine sediments, and the availability of P for biological uptake
from sediments depends on the distribution of sediment P within these fractions. For
example, apatite-P and aluminium–bound P are sinks of P and not readily available for
biological uptake (Paytan & McLaughlin 2007; Sinkko et al. 2011). Conversely, loosely
adsorbed P is readily available for uptake, while iron-bound P can become available when
Fe-compounds are reduced, and organic P can become available for uptake after
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mineralization (Sinkko et al. 2011). It is essential to measure more than just total P in
sediments, as different P fractions (with varying availabilities) can vary between different
sediments (Sinkko et al. 2011). However, in seagrass systems only a few studies have
examined P fractionation in sediments (Koch, Benz & Rudnick 2001; Holmer, Carta &
Andersen 2006), with the majority inferring availability from porewater P or total P in
sediments (Romero et al. 2006).
As described above, the extent of initial adsorption of inorganic P into the sedimentary
pool will largely depend on the mineralogy and clay content of the sediments. For
example, phosphate ions in solution strongly adsorbs to the surface of calcareous
sedimentary particles, limiting the availability of P for primary producers (Short 1987;
Duarte & Chiscano 1999). Phosphate can also be adsorbed onto Al and Fe oxide sediment
particles, but is more readily desorbed following reduction (Jordan et al. 2008). As such,
benthic producers growing in coastal systems dominated by calcareous sediments are
often considered to be P-limited (Short 1987; Walker et al. 1988; Marbà et al. 1996;
Hemminga et al. 1999). However, this generalization does not always hold true
(Erftemeijer et al. 1994; Ferdie & Fourqurean 2004; Fraser et al. 2012), suggesting that
other key processes, including the turnover of organic matter, may play a role in
maintaining P supply in some seagrass ecosystems.
Coastal sediments are not just a static reservoir of P, but are subject to biological and
geochemical processes that redistribute P in ecosystems (Benitez-Nelson 2000). The
recycling of P from sediments through plant uptake and then return of OM to the sediment
surface can be a significant input in the P budget of coastal ecosystems, often larger than
that contributed by allochthonous inputs from mangroves, rivers, or estuaries (Hemminga
et al. 1999; Ahlgren et al. 2006; 2011). Despite this, the majority of studies examining
marine P cycling have focused on open water systems and thus cycling in the water
column, with benthic cycling in sediments only recently receiving limited attention
(Stapel et al. 1996; Nielsen et al. 2007; Jilbert et al. 2011; Goldhammer et al. 2011).
However, benthic recycling can supply a substantial proportion of nutrients for primary
producers in coastal ecosystems, particularly in oligotrophic systems, and is likely
disproportionately important for nutrient cycling within these environments compared to
systems with higher phosphate concentrations (Fisher, Carlson & Barber 1982; Denis et
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Phosphorus cycling in seagrass sediments
al. 2001; Jahnke et al. 2005; Orth et al. 2006; Rauch & Denis 2008).
Organic matter and nutrient cycling in near coastal ecosystems
Highly productive seagrass meadows in oligotrophic conditions often grow in sediments
enriched with OM (Marbà et al. 1996; Kendrick et al. 2012). Seagrasses increase OM
concentrations in underlying sediments, primarily through deposition of dead plant
material as litter and increased sedimentation rates of plankton (Marbà et al. 2006). OM
contents in some seagrass sediments are as high as 100% (Fourqurean et al. 2012a). This
OM is often assumed to be a refractory pool (Kennedy et al. 2010), but the potential for
this OM to sustain productivity in seagrass ecosystems through mineralization has not
been examined in great detail. Nutrient release from the breakdown of OM is one of the
dominant sources of production, and is particularly important in sustaining productivity
in terrestrial oligotrophic ecosystems (Cleveland, Houlton & Smith 2013). For example,
the decomposition of OM supports primary productivity of tropical forests growing on
low nutrient soils (Reed et al. 2011), in freshwater lakes, streams and rivers (Gächter &
Meyer 1993), and in marine ecosystems (Leote et al. 2015). Globally, recycling of OM
accounts for 89% of the annual N demand and 98% of the annual P demand (Cleveland
et al. 2013). Nevertheless, understanding of the ecological impacts of OM in seagrass
sediments is a recognized knowledge gap (Romero et al. 2006). In oligotrophic
ecosystems, it is likely that the degradation of OM in seagrass sediments will be
particularly important from an ecological perspective by supplementing nutrient
availability.
Phosphorus availability and organic matter turnover
There have been remarkably few studies that have characterized sedimentary organic P
in seagrass sediments or the processes contributing to P turnover. The recycling of organic
compounds containing P is an important and complex component of the P cycle
(Grierson, Comerford & Jokela 1998; 1999; Grierson & Adams 2000; Turner, Frossard
& Baldwin 2005b). In marine sediments, organic P can comprise a large proportion of the
total P in sediments, and is a highly important component of the biogeochemical cycling
of P more generally (Gächter & Meyer 1993; Hayes, Richardson & Simpson 2000; Turner
et al. 2005a; Giesler et al. 2012). Traditionally, the organic P pool in both soils and
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sediments was considered to be refractory (Ahlgren et al. 2011). However, organic P is
comprised of a complex mix of known and unknown compounds (Turner, Newman &
Newman 2006; Reitzel et al. 2006; Turner 2008; Vincent, Turner & Tanner 2010; Turner
& Engelbrecht 2011). These compounds likely differ significantly in their availability for
plant uptake (Tarafdar & Claassen 1988; Gerke 2015). For example, phosphate
monoesters are of limited availability to plants as they are strongly associated with soil
minerals, while phosphate diesters or P lipids are more available for plant uptake due to
higher lability (Reitzel et al. 2006, Turner et al. 2011). Changes in the composition of the
soil organic P pool may lead to changes in productivity and species composition of
primary producers in terrestrial ecosystems (Turner 2008). We might expect that seagrass
communities would be similarly influenced by shifts in the amount of OM and its
interaction with sediment P fractions. Thus, information on both the amount and chemical
composition of sedimentary organic P is fundamental for understanding of P dynamics in
aquatic ecosystems (Turner et al. 2006; Zhang et al. 2009). Given low soluble reactive
phosphaotus (SRP) concentrations are common in marine waters overlying embayments
dominated by calcareous sediments, the recycling of P from the organic P pool may
represent the primary source of P to seagrass communities.
The biological mineralization of OM by microorganisms is the dominant pathway for
organic P to be broken down to inorganic phosphate available for plant uptake (Bünemann
et al. 2007; Achat, Bakker & Morel 2009). Biological mineralization is responsible for
up to 90% of plant P acquisition in terrestrial environments, and is generally more
significant in oligotrophic environments with low inputs of nutrients (van der Heijden,
Bardgett & van Straalen 2008). For example, tropical forests are systems with low P
inputs and highly weathered soils, where the dominant source of P for uptake by the large
pools of aboveground biomass comes from microbially-driven breakdown of OM (Reed
et al. 2011). Similarly, the jarrah forests of southwest Australia have been shown to be
highly dependent on microbial mineralization of organic P (Grierson & Adams 2000).
However, the importance of microbial mineralization for P supply in seagrass
ecosystems, including those of coastal Western Australia, is largely unknown.
Microorganisms may also compete with seagrass for P through the uptake of phosphate
and immobilisation into microbial biomass; this process can accentuate P limitation of
plants (Reed et al. 2011; Marschner, Crowley & Rengel 2011). Therefore,
6
Phosphorus cycling in seagrass sediments
microorganisms play a complex role within the P cycle as they have the potential to both
increase and decrease availabilities of P to higher plants. Competition between plants and
microbes for P is particularly strong in nutrient limited systems (van der Heijden et al.
2008). The balance between whether microbial activity is dominated by immobilization
or mineralization is dependent on environmental conditions like the relative availability
of other nutrients – particularly N – and the “quality” of of OM (in its simplest sense
characterised by the C:N or C:N:P ratio, (Hobbie 1996; van der Heijden et al. 2008).
Consequently, we might expect sediment microbes may act as a sink or source of P in
seagrass ecosystems depending on environmental conditions.
Microbial ecology – engines that drive biogeochemical cycles
Microorganisms have been described as a ‘black box’ in ecosystem ecology in
recognition of their central importance in regulation biogeochemical processes, yet there
is a profound lack of data regarding their quantitative impacts on ecosystem function
(Tiedje, Asuming-Brempong & Nüsslein 1999). This is particularly true in seagrass
ecosystems. There have only been a limited number of studies examining sediment
microbial communities in seagrass ecosystems. Seagrass meadows harbor very high
bacterial diversity in sediments (Garcia-Martínez et al. 2009), containing microorganisms
that have the capacity to influence ecosystem processes yet are not found in other
ecosystems. For example, novel diazotrophs are present in some seagrass sediments that
would influence N cycling (Bagwell et al. 2002; Hamisi et al. 2013). Microbial
communities adjacent to seagrass roots differ from bulk sediment communities in that
Epsilon- and Gammaproteobacteria increase and Deltaproteobacteria decrease (Jensen,
Kühl & Preimé 2007). Native bacterial communities are also critical for seagrass
transplant success, reducing transplant mortality (Milbrandt, Greenawalt-Boswell &
Sokoloff 2008), suggesting the interactions between microbes and seagrasses are
ecologically important. However, microbial communities in seagrass communities have
primarily been identified from a taxonomic perspective; and the functional roles of these
microbial communities have not been investigated. This gap in knowledge limits our
understanding of the significance of microbial communities on plant productivity,
diversity, and overall ecosystem function in seagrass meadows.
Microbial communities are highly impacted by changes in a range of environmental
7
Fraser (2016)
conditions. Where environmental shifts influence the functional capacity of the microbial
communities, there are likely to be shifts in ecosystem processes (Fierer, Leff & Adams
2012). For example, changes in substrate type (Jeffries et al. 2011a), N availability (Fierer
et al. 2011), and salinity (Hollister et al. 2010) drive shifts in microbial community
function that would likely influence biogeochemical processes and ecosystem functions.
Microbial communities mediate many decomposition processes through the production
of key enzymes, such as phosphatase (Burns & Dick 2002; Burns et al. 2013). The
production of these enzymes is encoded for by functional genes such as PhoA and PhoX
(Sebastian & Ammerman 2009). As such, the presence (or absence) of functional genes
within a microbial community has been used to assess their potential role in driving key
ecosystem processes (Zak, Blackwood & Waldrop 2006), and the loss of one microbial
functional group (e.g. N fixation) could lead to declines in plant productivity (van der
Heijden et al. 2008). Where unique functional genes are gained or lost from the microbial
community, the effects of these environmental drivers on ecosystem function would be
most apparent. In seagrass sediments, microbial communities shift in response to the
presence of seagrass and changes in season (James et al 2006). Again, the sediment
microbial community was only examined on a taxonomic level; the potential influence of
these microbial shifts on the function of the seagrass ecosystems has not been examined,
and we know very little of other environmental drivers in seagrass systems that may
influence microbial community composition. Salinity is a major factor that influences the
structure and functioning of coastal ecosystems. The majority of studies examining the
impacts of salinity on sediment microbial communities have examined transitions from
freshwater to brine to marine salinities (<5-35‰). For example, shifts in microbial
communities taxonomy and function in the Baltic Sea were most strongly associated with
salinity, including shifts in genes encoding for respiration that would affect C cycling
through the system (Dupont et al. 2014). Similar shifts under higher salinities (e.g.
hypersaline ecosystems) have received less attention. In hypersaline coastal wetlands, the
genes encoding for halotolerance and photosynthesis increased in sediment microbial
communities as salinities increased from 37‰ to >100‰ (Jeffries et al. 2011b). Such
shifts would also thus have influence on nutrient cycling and ecology across the entire
ecosystem.
In addition to salinity, nutrient availability gradients will also likely exert significant
8
Phosphorus cycling in seagrass sediments
impacts on microbial community composition in sediments (Fierer et al. 2011). Where
nutrients become more limiting, genes that promote efficient nutrient uptake and
recycling increase, given the microbial requirement for nutrients (Orchard & Webb
2009), altering nutrient cycling rates. Such changes in microbial communities are likely
to have the biggest impact in oligotrophic ecosystems where functional redundancy is low
(van der Heijden et al. 2008). For example, increased P limitation in microbes across
chronosequence soils shifted the functional capacity of the microbial community,
reducing decomposition rates and further intensified nutrient limitation (Wardle et al.
2004). Similar shifts in microbial communities along P availability gradients are likely to
influence ecosystem processes in seagrass ecosystems. Clearly, understanding the
functional capabilities of microbial communities by examining key genes is critical for
identifying microbial controls on biogeochemical cycling of P and relationships with
seagrass productivity.
Shark Bay: a model oligotrophic ecosystem
The research in this thesis is focused on Shark Bay, Western Australia (25°55`60°S,
113°32`32°E). Shark Bay is an extreme example of an oligotrophic coastal embayment,
with very low (0.02μM – 0.2 μM) concentrations of SRP in the water column (Atkinson,
1987; Fraser et al., 2012). Shark Bay is also an example of an inverse estuary (salinity
increases with distance from the ocean) with hypersaline reaches. The salinity gradient is
particularly pronounced across the eastern embayment of Shark Bay, rising from 35‰ in
the north to over 65‰ in the southern reaches of the Bay. This salinity gradient is
inversely correlated with water SRP concentrations in Shark Bay; P concentrations are
below detectable limits in hypersaline areas (Atkinson 1987; Fraser et al. 2012). Despite
the low P concentrations in the water column, Shark Bay contains productive seagrass
meadows that cover over 4000 km2 (~30%) of the embayment. The dominant seagrass
species are Amphibolis antarctica and Posidonia australis Hook.f. (Kendrick et al. 2012).
Both species are large-bodied temperate seagrasses capable of producing large root
systems that can take up nutrients from sediments (Pedersen, Paling & Walker 1997;
Evrard et al. 2005; Hovey, Cambridge & Kendrick 2012). These meadows support
diverse and commercially important fauna, and provide numerous key ecosystem services
in the Bay (Kendrick et al. 2012). Seagrasses are clearly foundation species in Shark Bay
9
Fraser (2016)
but the question remains as to how these meadows thrive in such a seemingly inhospitable
environment?
Shark Bay has been a focus of biogeochemists and ecosystem ecologists for several
decades as a marine ecosystem with unique and extreme biogeochemical conditions.
Shark Bay is historically important in marine biogeochemistry more broadly as it has
been used as an example of a P limited embayment in key biogeochemical papers (Smith
& Atkinson 1983; Smith 1984b; a). This early work proposed that P limitation of benthic
communities (including seagrass) increased along the salinity gradient due to the decrease
in water SRP and total sedimentary P (Atkinson 1987). This pattern of increasing P
limitation seems to be reflected in benthic nutrient content across the entire Bay
(~100km), with C:P ratios in seagrasses and other benthos increasing along with the
salinity gradient (Atkinson 1987; Burkholder, Fourqurean & Heithaus 2013). However,
there is considerable variation in C:N:P ratios across smaller spatial scales (10 km),
suggesting that factors other than salinity are driving P availability to seagrasses over
local scales in Shark Bay (Fraser et al. 2012). Shark Bay has relatively high
concentrations of OM in sediments (mean = 3.03%. Fourqurean et al. 2012a) and has
even been named as a carbon sequestration hotspot for coastal ecosystems globally
(Fourqurean et al. 2012b; a). Given the ecological importance of OM breakdown in
maintaining productivity of other terrestrial and aquatic ecosystems (described above),
we would expect that OM turnover is extremely important in supplementing nutrient
availability for seagrass meadows in Shark Bay.
Thesis aims and structure
The broad aim of this PhD thesis was to characterise the sources and ecological impacts
of organic matter inputs into seagrass sediments, with particular emphasis on
understanding the significance of OM decomposition and mineralization in determining
phosphorus availability in an oligotrophic seagrass ecosystem. In particular, the principle
objectives of this research were to investigate (i) if extreme climatic events may increase
the deposition of OM in sediments, (ii) the effect of seagrass-derived OM on sediment
biogeochemistry, microbial community composition, and seagrass physiology, (iii) if
organic P proportionally increased in sediments as salinity increased, and (iv) how
microbial utilization of OM changed along a gradient of increasing salinities and
decreasing SRP concentrations.
10
Phosphorus cycling in seagrass sediments
Chapter 2 and 3 examine the prospect and influence of OM deposition in oligotrophic
seagrass sediments, specifically increases in the deposition of seagrass detritus in
sediments (Figure 1.1). Chapter 2 documents widespread defoliation and dieback of the
dominant seagrass A. antarctica in Shark Bay following a marine heatwave and flood
event. These extreme climatic events lead to a large input of seagrass detritus into Shark
Bay sediments, representing an organic pulse that would influence the biogeochemical
conditions across this oligotrophic bay.
Figure 1.1 Conceptual model showing main questions in this thesis. 1) What are the impacts of extreme
climatic events on the seagrasses of Shark Bay? 2) How do inputs of seagrass-derived organic matter
influence the sediment biogeochemistry, microbial community structure, and seagrass seedling health? 3)
How does the fractionation of sedimentary P change along the salinity/phosphate gradient in Shark Bay?
What compounds dominate the organic P pool in seagrass sediments? 4) How does microbial community
structure and function change along the salinity/P availability gradient in Shark Bay sediments? Are
microbial communities adapted for low P concentrations at oligotrophic sites? 5) Does the microbial
turnover of organic P subsidise seagrass P requirements at sites with low soluble reactive phosphate
concentrations?
Chapter 3 examines the impacts of seagrass detritus in sediments using an experimental
tank system to control for other environmental factors. In this experiment, I investigated
how the addition of OM (as seagrass detritus), influenced sediment biogeochemistry and
microbial communities, and seagrass seedling health.
Chapters 4-6 focus on the links between OM decomposition and phosphorus
biogeochemistry in oligotrophic seagrass sediments. I characterised the different fractions
of sedimentary phosphorus across the salinity gradient in Shark Bay sediments, including
examining the organic P pool to the compound level using 31P-NMR (Chapter 4).
11
Fraser (2016)
In Chapter 5, I assessed the shift in microbial community structure and function along the
salinity/P availability gradient in Shark Bay sediments, and examined correlations to
changes in enzyme expression and the genes encoding for organic P metabolism along
the salinity/P availability gradient.
In Chapter 6, I describe attempts to use the stable oxygen isotope composition of
phosphates (δ18OPO4) as a tracer for biological cycling of P. My aim was to use this
technique to investigate the extent to which microbial turnover of organic P plays in
seagrass uptake of P across the salinity and P availability gradient. However, my attempts
to measure δ18OPO4 in seagrass and sediment samples were unsuccessful, despite
successfully processing and obtaining correct δ18OPO4 for international isotope standards.
Extremely low phosphate concentrations and high salinity (>50‰) prevented successful
extraction of pure silver phosphates and led to incorrect δ18OPO4 for Shark Bay seagrass
and sediment samples. As such, Chapter 6 is included here as a brief discussion of the
methodological challenges of applying this technique to the seagrasses and sediments of
Shark Bay, which provides a basis for future attempts for δ18OPO4 analysis in hypersaline
environments.
Finally, Chapter 7 provides a general discussion of the outcomes of this research, their
relevance to the understanding of organic matter dynamics and P cycling in the sediments
of Shark Bay, and possible avenues for further research.
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Fraser (2016)
Chapter 2: Extreme climate events lower resilience of
foundation seagrass at edge of biogeographical range
Matthew W. Fraser, Gary A. Kendrick, John Statton, Renae K. Hovey, Andrea Zavala
Perez, Diana I. Walker
Preamble: This paper was published in the Journal of Ecology (Fraser et al. 2014), and is
formatted to the requirements of the journal (including having a summary instead of a
traditional abstract). The work is primarily my own. GAK and DIW received the funding
for this project, and assisted me with experimental design. GAK, JS, RKH and AZP all
helped me with sample collection and processing. All co-authors provided feedback
during the writing of the manuscript. In this chapter, I document the widespread
defoliation and dieback of the dominant seagrass A. antarctica in Shark Bay following a
marine heatwave and flood event. This led to high inputs of seagrass-derived OM entering
sediments, with implications for later chapters.
Citation: Fraser MW, Kendrick GA, Statton J, Hovey RK, Zavala Perez A, Walker DI
(2014) “Extreme climate events lower resilience of foundation seagrass at edge of
biogeographical range.” Journal of Ecology 102:1528–1536. doi: 10.1111/13652745.12300
20
Phosphorus cycling in seagrass sediments
Summary
1. Extreme climatic events will dictate the response of ecosystems to climate change,
yet are understudied in marine ecosystems. The interaction of stressors from such
events has the potential to amplify negative impacts and drive ecosystems into
alternate states.
2. Here, we show a drastic response of a temperate seagrass species (Amphibolis
antarctica) in Shark Bay - a World Heritage Site in Western Australia at a
temperate-tropical transition zone - to two stressors driven by concurrent extreme
climatic events: a marine heatwave (Ningaloo Niña) and the Gascoyne Floods that
impacted the west coast of Australia in the austral summer of 2010-2011.
3. Widespread defoliation (leaf loss) of A. antarctica was observed in the months
following the extreme events, and were highest at sites affected by flooding
(Wooramel River Floods). We propose that the negative impact was magnified by
the synergistic interactions both stressors had on the carbon balance of the plant.
The elevated temperatures increased plant demand for carbon, which could not be
met through photosynthesis due to turbid floodwaters reducing light availability,
resulting in the plant having a negative carbon balance.
4. Recovery of leaf biomass was evident two years after the extreme events, , though
still 7-20% of historical averages. In contrast, below-ground biomass decreased
by an order of magnitude in the two years following the events. As below-ground
reserves underpin the tolerance of large seagrass species like A. antarctica to
disturbances, the declining trajectory of below-ground biomass will likely
manifest as a loss of resilience in A. antarctica to future disturbances.
5. Synthesis. Given the ecological importance of A. antarctica in Shark Bay as a
foundation species - accounting for 85% (~ 3700 km2) of the cover of seagrasses
in Shark Bay - predicted increases in the frequency and magnitude of similar
climatic events could have catastrophic implications for the future of this World
Heritage embayment. Where extreme climatic events overlap and cause multiple,
synergistic stressors to plant communities, ecological responses are likely to be
more extreme, particularly in ecosystems where foundation species exist near
upper thermal tolerance limits.
21
Fraser (2016)
Introduction
Climate change research is typically concerned with gradual climatic trends, yet extreme
climatic events will also dictate the response of ecosystems to climate change.
Understanding how ecosystems respond to extreme climatic events is therefore necessary
to predict how ecosystems and biodiversity will respond to climate change (Jentsch,
Kreyling & Beierkuhnlein 2007), especially as extreme events are predicted to become
more frequent and intense (Easterling et al. 2000). In particular, understanding the
response of foundational plant species to extreme climatic events is important, as this will
largely shape the ecological response at an ecosystem scale (Jentsch et al. 2007; Royer et
al. 2011). However, ecological responses to extreme climatic events can be variable and
hard to predict due to species- and system-specific responses (Beierkuhnlein et al. 2011;
Arnone et al. 2011; Smith 2011a). To date, most research has focused on the response of
terrestrial ecosystems to extreme climatic events, but marine ecosystems may also exhibit
extreme ecological responses to these events.
Temperature is unsurprisingly the main focus of climate change research in marine
ecosystems, and the relationship between extreme temperatures and physiological
performance is relatively well understood (Harley et al. 2006). Temperature, however, is
only one of a suite of stressors likely to impact marine ecosystems as a consequence of
extreme climatic events. Shifts in freshwater inputs, wave energies, and sediment inputs
are other stressors that could result from extreme climatic events in marine ecosystems
(Doney et al. 2012). These stressors are unlikely to act singly in natural ecosystems,
especially during extreme climatic events. Instead, multiple stressors can result from one
extreme climatic, the overlap of two or more extreme climatic events, or in systems that
have pre-existing stressors unrelated to extreme events. When these multiple stressors
interact, the result can be synergistic where the combined effect is greater than the sum
of individual stressors (Darling & Côté 2008). These combined events have the potential
to evoke extreme ecological responses (Smith 2011a), reducing resilience or even push
ecosystems into alternate states. Chesapeake Bay in the US provides a classic example of
a system being pushed into an alternate state as a result of synergistic stressors (Orth &
Moore 1983). Accumulated loss of oyster beds, as well as the loss of other buffers such
as seagrass, wetlands and forests weakened the bay’s resilience and left it vulnerable to
torrential runoff brought by an extreme climatic event - Tropical Storm Agnes - in 1972
22
Phosphorus cycling in seagrass sediments
(Orth & Moore 1983). Multiple stressors are particularly concerning for management as
the combined effect are notoriously difficult to predict, yet have the potential to erode
resilience and lower the threshold for sudden phase shifts in ecosystems (Harley et al.
2006; Hawkins et al. 2008).
The ability of a system to retain essential functions, structure, identity and feedbacks
depends on multiple attributes such as the latitude (the maximum amount a system can
change before losing the ability to recover), resistance (the ease or difficulty of changing
the system) and precariousness (how close the current state of the system is to a limit or
threshold) (Walker et al. 2004). The final determining aspect of resilience is how the
above three attributes are influenced by states and dynamics of the system at scales above
and below the scale of interest (panarchy). As such, not all organisms or ecosystems are
equally susceptible to state shifts driven by climate related stressors. In ecosystems where
foundation species are at risk from multiple stressors, strong panarchial connections mean
that any associated mortality has the potential for catastrophic ecosystem-level effects.
Foundation species, like keystone predators (Paine 1966), have a disproportionately large
impact on the ecosystems they grow in, with significant cross-scale interactions creating
locally stable conditions that support other species and ecosystem processes (Ellison et
al. 2005). Loss of foundation species can lead to bottom-up impacts that permeate through
the trophic structure of the ecosystem. For example, loss of foundational tree species in
North American forests has decreased biodiversity and altered ecosystem processes such
as decomposition and nutrient cycling (Ellison et al. 2005). Ecosystems where foundation
species exist close to physiological thresholds (e.g. the edge of their biogeographical
range) increase the precariousness of the current state of the system, making them
particularly vulnerable to multiple stress events, and therefore escalate the potential for
catastrophic loss of ecosystem function and services.
Shark Bay is a World Heritage listed marine embayment with a foundational plant species
at the edge of its biogeographic range. The temperate seagrass Amphibolis antarctica is
the dominant primary producer in Shark Bay, covering approximately 3700 km2 (~85%
of seagrass covered area) of Shark Bay (Walker, Kendrick & McComb 1988),
undoubtedly making it the foundation species. One of the key environmental values that
this foundation species supports, and that the World Heritage status aims to preserve, is
23
Fraser (2016)
high marine biodiversity, including significant populations of dugongs, turtles, and tiger
sharks (Kendrick et al. 2012). Fish and dugongs forage on A. antarctica (Wirsing,
Heithaus & Dill 2007; Burkholder et al. 2013) and further evidence suggests that it is
indirectly responsible for the characteristic salinity gradient in the Bay, reducing water
flow into the upper reaches of the embayment allowing for the development of the world’s
most abundant population of stromatolites – mushroom shaped mounds made from layers
of cyanobacteria and sediment that form reefs and that are modern analogues for some of
Earth’s earliest life - in the hypersaline areas of Shark Bay (Logan, Cebulski & Logan
1970; Jahnert & Collins 2011). These seagrass meadows also sequester some of the
highest below-ground carbon stores seen in seagrass meadows worldwide (Fourqurean et
al. 2012a). Amphibolis antarctica in Shark Bay are near the upper limit of the species
thermal tolerance. As such, A. antarctica growing in Shark Bay is in a more precarious
position when exposed to above-average temperatures or other stressors (Walker &
Cambridge 1995). Given the panarchial connections of foundation species, any loss of A.
antarctica in Shark Bay may have severe consequences for the entire ecosystem and drive
it to an alternate state.
Here, we show a potential decrease of resilience in Amphibolis antarctica meadows in
Shark Bay, with short term (0-6 months) leaf loss and long term (2 years) loss of belowground biomass. We link this loss of resilience to the synergistic stressors of an extreme
temperature event (a marine heatwave) and extreme rainfall events (the Gascoyne Floods)
that impacted the west Australian coast in the austral summer of 2010-2011. We then
discuss the potential future trajectories of A. antarctica in Shark Bay given projected
increases in extreme climatic events, and discuss potential ecosystem level implications
for the Shark Bay World Heritage Area.
24
Phosphorus cycling in seagrass sediments
Methods
Study site
The study was conducted at the eastern embayment of Shark Bay, Western Australia (25°
55' 60 S, 113° 32' 32 E), a 13000 km2 marine embayment (Fig. 2.1a,b). The climate of
Shark Bay is hot and dry, with annual evaporation (2000mm) being ten times annual
precipitation (200mm). There is a strong salinity gradient in the eastern embayment (36
to > 65‰), driven by the high evaporation rates and restricted circulation with Indian
Ocean water (Kendrick et al. 2012). Water temperatures in Shark Bay range from 17°C
in winter to 26°C in summer (Walker et al. 1988). Shark Bay is oligotrophic, with
phosphorus concentrations below detectable limits in the water column (Fraser et al.
2012). Shark Bay contains some of the most extensive seagrass meadows in the world,
and contains up to 12 different species of seagrass (Walker et al. 1988), making it a global
hotspot of seagrass diversity. Shark Bay is dominated by monospecific meadows of A.
antarctica that cover 3700 km2 of the Bay. Posidonia australis and Halodule uninervis
(Forsk.) Aschers. also occur in large monospecific meadows in the Bay (Walker et al.
1988). The Wooramel River is the only river that discharges directly into Shark Bay,
though only flows during periods of cyclonic events and winter storms. Data on daily sea
surface temperatures were downloaded from NOAA’s ¼ degree gridded Advanced Very
High-Resolution Radiometer Optimally Interpolated Sea Surface Temperature (Reynolds
et al. 2007). Mean monthly sea surface temperature was then calculated from this data.
Data on monthly discharge from Wooramel River were downloaded from the Department
of
Water,
Western
Australia
(http://kumina.water.wa.gov.au/waterinformation/wir/reports/publish/703002/g02.htm).
Spatial extent of Amphibolis defoliation
Twenty-three sites across the eastern embayment were visited in March 2011 and 11 sites
were visited in September 2011, incorporating a range of distances to the Wooramel River
(6.5 km to 29.9 km; Fig. 2.1c). Water depth was similar across all sites (1-3 m deep). Five
sites within 15 km of the River were visited once in March 2011, once in September 2011,
and once in March 2013, to investigate longer-term recovery in sites where greatest
defoliation was recorded.
Amphibolis antarctica was collected at each site with a 25 cm diameter core (18 cm deep)
25
Fraser (2016)
from the centre of each of five 0.25 m2 plots that allowed for both above and belowground biomass to be sampled in entirety. Plots were randomly chosen by dropping a
quadrat from the boat, and flipping the quadrat three times for each individual plot. The
entire contents of the core (including below-ground seagrass biomass and sediment) were
extracted intact, washed free of sediment and sorted into species. All samples were stored
on ice in the field, and then frozen immediately for later processing. Seagrass samples
were thawed and epiphytes and sediments removed from seagrass leaves by gentle
scraping with a scalpel under flowing deionised water. Amphibolis antarctica samples
were separated into leaves, shoots, rhizomes, and roots. All dead seagrass biomass
(necrotic) was removed prior to weighing. Seagrass was oven dried at 60°C until constant
weight, and weighed to estimate biomass.
Data analysis
Leaf biomass was regressed against distance from the Wooramel River. Differences in
leaf and below-ground biomass were determined using ANOVA. A generalized additive
model was used to determine the relationship between seagrass biomass and distance
from Wooramel River. Data were transformed where assumptions of heterogeneity were
not met. R (version 2.13.1) was used for all statistical analysis.
26
Phosphorus cycling in seagrass sediments
Fig. 2.1 Location of the study area. (a) Location of Shark Bay. (b) Location of the study site in the eastern
embayment of Shark Bay. (c) Location of individual sites within the study area. White and black boxes
show sites that were visited in March and September 2011 to determine the effect of proximity to the
Wooramel River on seagrass biomass. Black boxes show sites that were visited in March 2011, September
2011, and March 2013 to determine seagrass biomass and potential recovery after the extreme climatic
events. Sampling protocols were consistent across all sites and study periods.
27
Fraser (2016)
Results
Extreme temperature and rainfall
In 2011 an anomalous rise in sea surface temperatures (SST) was observed off the west
coast of Australia, with peak SST 5°C warmer than normal (Feng et al. 2013). This marine
heatwave, dubbed the ‘Ningaloo Niña’, was driven by intense Leeuwin current flows, an
extraordinary La Niña, and multi-decadal trends in the Pacific Ocean (Feng et al. 2013).
The combination of these events was unprecedented, and can therefore be considered an
extreme climatic event (Smith 2011b).
In Shark Bay, the mean monthly SST reached a maximum of 29°C in February 2011, and
was above 26°C in January, February, March, and April 2011 (Fig. 2.2a). Mean monthly
SST was above 2006-2010 averages in January (2.8ºC), February (3.8ºC), March (2.0ºC),
and April 2011 (1.8ºC).
A large rainfall event over the Gascoyne region of Western Australia occurred on the 15th
December 2010. This rainfall resulted in high river flow of the Wooramel River from
December 2010 – February 2011 (Fig. 2.2b), reaching a maximum of 300 Gl discharged
in February 2011. This prolonged flood event was extreme for the Wooramel River; with
river discharge higher in December 2010 (270 Gl) and February 2011 (300 Gl) than any
other month in the previous decade (Fig. 2.2b). The floodwaters brought significant
volumes of suspended terrestrial sediments, with flood plumes apparent 10-15 km from
the River mouth (Figs 2.3a,b). In sites within 10 km from the river mouth, up to 10 cm of
sediment was deposited on seagrass meadows (M.W. Fraser, unpublished data).
28
Phosphorus cycling in seagrass sediments
Average temperature (°C)
30
(a)
Marine Heatwave
25
20
01/2004
01/2006
01/2008
01/2010
01/2012
Date
Discharge volume (gigalitres)
300
Gascoyne Floods
(b)
200
100
0
01/2004
01/2006
01/2008
01/2010
01/2012
Date
Fig. 2.2 (a) Mean monthly sea surface temperature (°C) in Shark Bay, Western Australia between January
2004 and January 2012. Black outline denotes 2010/2011 ‘marine heatwave’. Data were downloaded from
NASA’s MODIS Aqua 4km dataset. (b) Monthly discharge (Gigalitres) from the Wooramel River, between
January 2004 and January 2012. Data were downloaded from the Western Australian Department of
Water’s Wooramel River station.
Response of seagrasses to flooding and temperature events
Amphibolis antarctica defoliation (leaf loss) was apparent in meadows in March 2011
(Figs 2.3c,d). Mean A. antarctica leaf biomass ranged from 9.2 g m-2 to 342.3 g m-2 in
29
Fraser (2016)
March 2011. Leaf biomass was significantly higher as distance from the Wooramel Delta
increased in March (R2 = 0.694, p < 0.0001, n=23). Leaf biomass only increased above
100 g m-2 at sites further than 14 km from the Wooramel River mouth in March 2011
(Fig. 2.4a), and was always greater at sites further than 14 km from the River mouth. In
September 2011, six months after the flood and heatwave events, this pattern in leaf
biomass persisted (Fig. 2.4b). Leaf biomass was significantly higher as distance from the
Wooramel River increased up to 15 km, after which leaf biomass stayed steady at around
180 g m-2 (R2 = 0.694, p < 0.0001, n=23). In the only other seagrass species encountered,
Halodule uninervis, leaf biomass was not significantly affected by distance to the
Wooramel River (R2 = 0.01, p =0.63, n=22).
Fig. 2.3 (a) Satellite imagery showing study area before the 2010/11 Wooramel River floods. (b) Satellite
imagery showing study area 1 month after the 2010/11 Wooramel River floods (courtesy of Curtin
University: January 08 2011 RSSRG). Note the flood plume extending out from the coastline. (c) Healthy
Amphibolis antarctica meadow, showing plants with numerous green leaves. (d) Defoliated Amphibolis
antarctica in Shark Bay, showing dead plants with all leaf material lost. Photo credit: Gary Kendrick.
30
Phosphorus cycling in seagrass sediments
Recovery of seagrass meadows
Mean leaf biomass at sites close to the river increased from 8.4 g m-2 to 39.9 g m-2 between
March and September 2011 (Fig. 2.5a). By March 2013 - two years after the flood and
heatwave events - mean leaf biomass at sites closer to the river had increased to an average
of 45.4 g m-2. Leaf biomass was higher at sites further from the Wooramel River during
all sampling periods, increasing from 52.9 g m-2 to 68.6 g m-2 between March and
September 2011. Mean leaf biomass then increased to 116.1 g m-2 two years after the
flood and heatwave event, still far lower than the mean leaf biomass measured in a
comparable study using similar methodologies carried out in the same general location
(e.g. at similar salinities on the Wooramel Delta) in the 1980s (~600 ± 200 g m -2; (Walker
1985)).
Mean below-ground biomass followed the opposite pattern to leaf biomass, decreasing
between March 2011 and March 2013 across all sites (Fig. 2.5b). In sites closer to the
Wooramel River, below-ground biomass decreased by an order of magnitude between
March 2011 (208.4 g m-2) and March 2013 (20.8 g m-2). At sites further from the river,
mean below-ground biomass decreased across all three sampling periods, falling from
284 g m-2 in March 2011 to 51 g m-2 in March 2013.
31
Fraser (2016)
●
Leaf biomass (g m−2 )
(a)
●
●
●
● ●
● ●
100
●
●
●
●
●
●
●●
●
●
●●
10
●
●
●
10
15
20
25
30
Leaf biomass (g m−2 )
(b)
●
● ●
●
100
●●
●
●
●
●
10
●
●
10
15
20
25
Distance from Wooramel River (km)
Fig. 2.4 Scatterplots comparing Amphibolis antarctica leaf biomass collected during (a) March 2011 and
(b) September 2011 with distance from the Wooramel River. The blue lines represent generalised additive
models fitted to the data, grey outline shows 95% confidence interval. Note the logarithmic scale on the y
axis.
Discussion
In Shark Bay, Amphibolis antarctica is the most abundant seagrass species, occupying
3700 km2 (~85% of seagrass covered area) of the shallow subtidal (1-5m depth, Walker
et al. 1988), and creates locally stable conditions that support faunal communities and
ecological processes (Wirsing et al. 2007; Heithaus et al. 2007), so any negative impacts
of global climate change on A. antarctica would have far-reaching effects across the
Shark Bay ecosystem. We have found that A. antarctica underwent defoliation after an
32
Phosphorus cycling in seagrass sediments
extreme climatic event (the marine heatwave). Where the marine heatwave overlapped
with the turbid floodwaters from the Wooramel River (the result of another extreme
event), there was a more extreme defoliation response (Figs 2.4 and 2.5). We propose that
this was a consequence of the interaction between the elevated temperatures and higher
turbidity that reduced light availability. Two years after the extreme events, there has been
an increase in leaf biomass, suggesting A. antarctica is recovering, but below-ground
biomass has declined. As below-ground reserves underpin the tolerance of seagrasses to
perturbations (Alcoverro et al. 1999), the declining trajectory of below-ground biomass
will likely manifest as a loss of resilience in A. antarctica. Thus, the reduced capacity of
the systems foundation species to withstand future perturbations could potentially drive
the Shark Bay ecosystem to an alternate state. More generally, ecological responses to
extreme climatic events can be variable and hard to predict due to species- and systemspecific responses (Smith 2011a). Ecosystems with foundation species growing near
upper thermal tolerance limits (like A. antarctica in Shark Bay) will be more likely to
show extreme ecological responses after extreme climatic events.
33
Fraser (2016)
Leaf biomass (g m−2 )
(a)
100
50
0
Belowground biomass (g m−2 )
0 months
6 months
2 years
6 months
2 years
(b)
300
200
100
0
0 months
Time after heatwave
Fig. 2.5 Mean Amphibolis antarctica (a) leaf biomass and (b) below-ground biomass (rhizome and root
biomass) collected in Shark Bay in March 2011 (0 months), September 2011 (6 months), and March 2013
(2 years after the heatwave/flood event). Dark grey bars denote sites within 12 km of the Wooramel River,
and light grey bars denote sites further than 12 km from the mouth of the Wooramel River. Error bars
indicate ± 1 s.e. Note that mean leaf biomass measured in the 1980s (600 g m-2; Walker 1985) would far
exceed the scale of leaf biomass measured in this study.
While A. antarctica near the Wooramel River suffered extensive loss of biomass after
exposure to both increased temperatures and highly turbid floodwaters, biomass was
sustained away from the influence of the turbid waters where sufficient light was able to
reach the benthos. This was common across Shark Bay, with meadows in turbid sites off
Monkey Mia experiencing highest mortality of seagrasses, while sites characterised by
clearer waters were less affected (J.A. Thomson, unpublished data). Large seagrass
34
Phosphorus cycling in seagrass sediments
species like A. antarctica have high light requirements because of their need to balance
photosynthetic carbon production with the large respiratory demand from nonphotosynthetic below-ground tissues (Hemminga, Marbà & Stapel 1999). Increases in
water temperature cause greater respiratory demand for carbon in seagrasses (Ralph et al.
2007), and can also inhibit photochemical efficiency and damage photosynthetic
pathways (Campbell, McKenzie & Kerville 2006). The interplay between these stressors
has also been noted for A. antarctica’s congeneric Amphibolis griffithii in South-Western
Australia (Lavery et al. 2009). The significance of catastrophic loss in biomass presented
here is that these episodic climatic events are predicted to occur more frequently, eroding
the resilience of a foundation species that is already close to its thermal threshold. As
mentioned previously, diminished resilience of the foundation species leading to
significant loss will likely have a cascading effect on the system, with potential for shifts
to alternate states via ecological extinction of entire trophic levels (Ellison et al. 2005;
Dobson et al. 2006). For a system like Shark Bay where the current ecosystem services
such as high biodiversity exist due to the presence of A. antarctica, any loss of resilience
in this species poses a threat to its ecological value and its ability to meet World Heritage
status criterion.
Below-ground biomass declined for two years, suggesting that the flood and heatwave
events may have a longer-lasting, legacy impact on A. antarctica stands, or that this
reallocation of biomass may be a strategy for survival within this species. Seagrass use
below-ground tissues as stores to initiate recovery or persist through unfavourable
conditions (Alcoverro, Manzanera & Romero 2001). Reallocation of biomass from
belowground compartments to leaves is a common plant response that contributes to
adaptive capacity to changing environmental conditions (McConnaughay & Coleman
1999; Weiner 2004). Despite A. antarctica displaying this plasticity, the lack of elasticity
(slow recovery time) may change the dimensions of basins of attraction within the
resilience landscape, reducing the latitude and resistance to future perturbations.
Recovery of leaf biomass at sites to levels measured in the 1980s (Walker 1985) would
take 37-38 years at sites close to the river, and 17-18 years at sites further from the river,
assuming a linear relationship derived from our data (Fig. 2.5a; y = 2.63x + 52.83 for
sites within 12 km of the Wooramel River and y = 1.26x - 18.71 for sites further than 12
km from the river). However, climatic models predict an increase in intensity and
35
Fraser (2016)
frequency of extreme events in Shark Bay (Alexander & Arblaster 2009). Even if
decreased belowground biomass aids survival of A. antarctica by reducing respiratory C
demand, should the frequency of extreme events exceeds recovery times of A. antarctica,
then the resilience of A. antarctica to disturbances is likely to be eroded. Existing A.
antarctica stands may persist in the short term, but in a condition where they are at more
risk to future stressors such as increased temperatures, reduced water quality, disease and
physical disturbance.
The flood and heatwave events have led to two responses in A. antarctica – a short term,
highly visible reduction in leaf biomass; and a long term, less visible reduction in belowground biomass. This presents a clear message to marine conservationists and managers
in Shark Bay and elsewhere; not to rely solely upon traditional above-ground metrics to
capture ecosystem health and functioning. Management strategies that mitigate local
stressors can promote resilience of benthic communities to inevitable disturbances such
as increased temperatures (Russell et al. 2009). For example, recovery of coral reefs after
bleaching events is quicker at sites without chronic human pressures (Carilli et al. 2009;
Gilmour et al. 2013), leading to recommendations that local pressures such as overfishing
and pollution should be reduced to aid resilience of corals to climate change (Thompson
& Dolman 2010; Ateweberhan, Feary & Keshavmurthy 2013). The Wooramel catchment
is dominated by pastoral farms with high rates of ungulate grazing that have led to
widespread sediment erosion (Curry et al. 1994), which contribute to the loss of terrestrial
vegetation and increase of sediment inputs during flood events (McKergow et al. 2005).
The reduction of sediment inputs through catchment management would help increase
long term survival prospects of A. antarctica growing on the Wooramel Delta.
As A. antarctica is a foundation species in Shark Bay, any loss would be expected to have
bottom up influences across the ecosystem if it is not replaced by a functionally equivalent
species. Halodule uninervis - another seagrass species – did not appear to suffer similar
dieback to A. antarctica following the heatwave and flood events, though was slightly
less likely to be observed in 2012 compared to 2007-2009 (Thomson et al. 2014).
Halodule uninervis is a tropical seagrass species with a life history strategy well suited to
disturbance regimes that include low light availabilities and high temperatures
(Carruthers et al. 2002). Halodule uninervis relies on a sedimentary seed bank to persist
36
Phosphorus cycling in seagrass sediments
through periods of light deprivation, unlike A. antarctica that persists through
underground biomass stores. As such, H. uninervis may replace A. antarctica at sites that
are marginal for A. antarctica growth, such as the Wooramel Delta. However, these two
species are not functionally equivalent, and loss of A. antarctica could lead to significant
ecosystem-scale effects due to the ecological extinction of processes including those that
maintain niche habitats for species of high conservation priority. Indeed, even small
changes in the abundance of foundation species can potentially lead to feedbacks that
cause ecosystem wide effects (Byers et al. 2006). Loss of A. antarctica could directly
lead to loss of forage habitat for species of conservation concern including turtles and the
world’s largest population of dugongs (Wirsing et al. 2007; Heithaus et al. 2007). The
flow of important ecosystem services such as sediment trapping or carbon burial (Jahnert
& Collins 2011; Fourqurean et al. 2012b) will also be significantly impacted, leading to
significant ecological, social, and economical costs. Dead seagrass biomass would also
enter detrital pathways (Walker, Kendrick & McComb 2006), altering biogeochemical
cycling across this oligotrophic embayment. On a regional scale, loss of A. antarctica
may also lead to a biogeographical range contraction for this endemic seagrass. Though
care must be taken when interpreting changes resulting from a single extreme event due
to the implicit lack of replication, the documented loss of a foundation species will likely
have ecosystem scale consequences in Shark Bay.
State shifts can occur rapidly when drivers interact to alter competition between species
(Suding & Hobbs 2009). These abrupt changes in the structure of the system tend to be
maintained by feedback processes, making it difficult to return to the previous state
(Mayer & Rieterk 2004). Shifts in marine communities from domination by temperate to
tropical species has been noted in other marine plants and animals along the Western
Australian coastline as a result of the 2010/11 marine heatwave (Wernberg et al. 2012),
and loss of seagrass following marine heatwaves has been noted in the Mediterranean
(Marbà & Duarte 2009). Poleward movement of A. antarctica to escape future warming
trends would be severely restricted in Shark Bay given its geomorphology, and may lead
to a reduction in connectivity of this species along the West Australian coastline (Opdam
& Wascher 2004). Loss of A. antarctica from Shark Bay would result in dramatic changes
to the food web and structure of the shallow sills and banks that are characteristic of the
Shark Bay ecosystem, potentially leading to catastrophic collapse of the seagrass
37
Fraser (2016)
dominated ecosystem. Large scale seagrass loss (>1000 km2) in other seagrass dominated
systems such as Chesapeake Bay and Cockburn Sound have resulted in significant
impacts on benthic metabolism due to large quantities of organic matter entering detrital
pathways, and subsequent changes in the physical environment can preclude future
recolonization (Kendrick et al. 2002; Kemp et al. 2005). Clearly, multiple stressors
resulting from extreme events must be considered alongside gradual warming trends to
accurately predict, and mitigate against, the effects of climate change in marine
ecosystems.
Acknowledgments
A NHT-II Caring for Country Project grant coordinated by Western Australian Marine
Science Institute, and awarded to D.I.W. and G.A.K., funded this research. Additional
funding for M.W.F. was received through the Ray Hart Memorial Scholarship, an
Australian Geographic Bayerboost Scholarship, and Ph.D. funding from the School of
Plant Biology, UWA. Logistical support was provided by the Shark Bay Ecosystem
Research Project under the direction of Michael Heithaus, funded by NSF (OCE0745606). We thank Yasha Hetzel for assistance in obtaining downloading SST data from
NOAA, and two anonymous reviewers for comments that contributed greatly to the
manuscript.
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Phosphorus cycling in seagrass sediments
Chapter 3: Seagrass derived organic matter influences
biogeochemistry, microbial communities, and seedling biomass
partitioning in seagrass sediments
M.W Fraser, J. Statton, R.K. Hovey, B. Laverock, G.A. Kendrick
Preamble: This paper was published in Plant & Soil (Fraser et al. 2015), and is formatted
to the requirements of the journal. The work is primarily my own. GAK and JS received
funding for the project. MWF, JS, and RKH contributed to upkeep and maintenance of
tank system. BL and MWF led microbial analysis, MWF and JS led seagrass analysis,
MWF carried out biogeochemical analysis. All co-authors contributed to experimental
design and provided feedback during the writing of this manuscript. In this chapter, I
examine the impacts of seagrass-derived OM on belowground biogeochemistry,
microbial communities, and seagrass seedling physiology in a controlled tank
environment. This provides context for the potential belowground effects of the dieback
of seagrass discussed in the previous chapter, as well as leading into the further
investigation of OM as a phosphorus source examined in subsequent chapters.
Citation: Fraser MW, Statton J, Hovey RK, Laverock B, & Kendrick GA. (2015)
“Seagrass derived organic matter influences biogeochemistry, microbial communities,
and seedling biomass partitioning in seagrass sediments” Plant and Soil.
doi:10.1007/s11104-015-2721-0
43
Fraser (2016)
Abstract
Aims Seedling establishment is a crucial life history stage in seagrasses, yet factors that
affect seedling health are poorly characterized. We investigated if organic matter (OM)
additions to sediments provided nutritional benefits for seagrass seedlings through
microbial degradation.
Methods We tested the effects of sedimentary OM additions on Posidonia australis
seedlings growing in tank cultures. We focused on sediment biogeochemical processes
and microbial communities that may impact seedling growth and physiology.
Results Enrichment of sediments with OM changed microbial community composition
(DNA-ARISA) and a significant increase in hydrolytic enzyme expression. Total seedling
biomass did not differ between OM treatments, but above:belowground biomass
increased with OM enrichment. Nitrogen and phosphorus concentration of seagrass
leaves was lower with increasing OM.
Conclusions Seagrass derived OM has been considered a refractory store of carbon, yet
here we show its deposition into sediments significantly alters belowground conditions.
Remineralization of the OM changes both physical and chemical nature of sediments that
leads to greater biochemical activity, change in microbial communities and greater
investment into above ground photosynthetic biomass. The presence of OM may assist
seagrass seedling survival during early development by enhancing root branching and
stability in sediments, but is unlikely to provide nutritional benefits.
Keywords: bacteria, decomposition, enzymes, Posidonia australis, roots, seeds
44
Phosphorus cycling in seagrass sediments
Introduction
Inorganic nutrient concentrations are often below concentrations required to maintain
primary productivity in marine sandy habitats where seagrasses dominate (McGlathery et
al. 2001). The microbial breakdown of organic matter (OM) supplements primary
productivity, and is especially important in low nutrient ecosystems such as tropical
forests (Reed et al. 2011). Similarly, OM buried in marine sediments may represent an
important source of nutrients to seagrasses (López et al. 1995; Evrard et al. 2005;
Kilminster et al. 2006). The role of microorganisms in OM breakdown may be
particularly important in seagrass sediments as these sediments contain high
concentrations of seagrass derived OM (Kennedy et al. 2010). While this belowground
pool has been widely investigated for its role in carbon (C) sequestration (Fourqurean et
al. 2012), the ecological implications of these sediments have not been examined in great
detail. Seagrass derived OM consists of a range of organic molecules that differ
significantly in their biological availability, but most forms of OM are refractory and not
readily available for direct uptake by seagrasses (Danovaro 1996). Seagrass therefore rely
on the metabolic capabilities of their sedimentary microbes to convert complex molecules
to more labile forms of dissolved OM and nutrients that are then available for uptake by
seagrasses. Conversely, microbes may also reduce nutrient availability through
immobilisation of available nutrients into microbial biomass.
Microbes primarily liberate nutrients from OM through the release of extracellular
enzymes. Extracellular enzymes play a crucial role in all biogeochemical cycles by being
proximate agents in crucial processes such as decomposition (Burns and Dick 2002;
Burns et al. 2013). Enzymes play a complex role in the degradation of plant litter, with
numerous enzymes needed to hydrolyse OM at different stages. Plant litter comprises a
high proportion of recalcitrant C rich macromolecules such as cellulose, lignin, or
hemicellulose. These macromolecules are broken down through the combined enzyme
activities of a diverse range of microorganisms (Burns and Dick 2002; Burns et al. 2013).
For example, β-glucosidase breaks down cellulose, while α-glucosidase hydrolyses starch
(Dick 2011). Consequently, the release of nutrients via microbial enzyme production is
partly dependent on the type and quantity of plant litter (Burns et al. 2013). For instance,
expression of β -glucosidase and acid phosphatase (associated with C and phosphorus (P)
45
Fraser (2016)
release from OM) in soils was higher in soils amended with grass litter relative to soils
without litter addition (Dornbush 2007). Distinct shifts in the microbial community
composition are also apparent during decomposition (Sinsabaugh and Follstad Shah
2012). These shifts can be directly (C availability, chemical composition of litter) or
indirectly (redox, pH) related to the presence of OM. The addition of OM to sediments
should therefore result in distinct belowground shifts in microbial communities and
enzyme production.
The successful establishment of seedlings is a crucial life history stage in
seagrasses. Recolonization through seedling establishment may be particularly important
in large degraded areas with high rates of seagrass mortality that limits recovery through
asexual propagation. Successful recovery would therefore likely depend on new recruits
entering from adjacent populations; with several biotic and abiotic bottlenecks to
successful settlement and growth of seedlings (Orth et al. 2005; Rivers et al. 2011).
However, ecological factors influencing seagrass seedling survival post settlement have
received limited attention, despite their importance for maintaining populations
(Kendrick et al. 2012). In particular, the nutrient requirements of recently settled seedlings
are poorly understood. Additions of inorganic nutrients (N and P) did not increase
seedling growth or productivity (Statton et al. 2014). Instead, the release of nutrients from
decomposing OM may be an important source of nutrients to P. australis seedlings.
Posidonia australis seedlings grown in tanks with additions of OM showed increased root
growth and productivity compared to those grown in control sediments (Statton et al.
2013). However, the belowground processes that increase this root growth are not known.
Here, we tested the effects of OM additions on P. australis seedling growth in
tank cultures. We also investigated sediment biogeochemical processes and microbial
communities that could explain seagrass responses to OM additions. We hypothesised
that (i) enzymes involved in C and P cycling would have higher concentrations at higher
OM concentrations, (ii) microbial communities would be significantly different in pots
with OM compared to pots without OM, and (iii) seagrass seedlings would display higher
leaf, root and rhizome biomass and nutrient (N and P) concentrations in sediments with
higher concentrations of OM.
46
Phosphorus cycling in seagrass sediments
Methods
Seedling collection
Posidonia australis fruit were collected from Woodman Point (32º08’10”S,
115º44’31”E) during November 2012. Fruit were transported in seawater back to the
Fisheries WA Marine Research Facility at Hillarys, Western Australia. Fruit were left to
dehisce naturally over 3 days in a holding tank containing constant flow-filtered seawater.
Dehiscence of fruit resulted in the negatively buoyant P. australis seeds (with emerging
radicle and prophyll) collecting in numbers on the bottom of the tank. We assessed seeds
for viability based on the presence of an intact seed, developing prophyll, and radicle.
Experimental design
The effect of OM addition on seagrass seedling growth and sediment biogeochemistry
was investigated using three tanks, each containing eight replicates of six organic matter
treatments. Tanks received a 12h/12h light regime supplied by artificial metal halide
lamps positioned above each tank (250 μmol photons m-2 s-1 or 10.8 mols photons d-1),
with constantly replaced seawater filtered to 0.2μm. Replicate seedlings were planted into
square pots (70 x 70 mm, 300 mm height) containing one of the six OM treatments (0,
0.5, 1, 1.5, 2, 2.5% sediment dry weight). Given the quantitative nature of our treatments,
we used a ‘replicated regression’ design to increase the power and flexibility of our
experiment (Cottingham et al. 2005). These OM concentrations represented levels
common in natural seagrass sediments, and are comparable to previous P. australis
aquaculture experiments (Statton et al. 2013). Sediment was collected from Woodman
Point beach. This sediment was primarily calcareous, and contained negligible quantities
of OM (<0.1% dry wt.) Sediment was dried at 60ºC for 72 hours prior to addition to pots.
Organic matter was sourced from Woodman Point beach wrack and consisted primarily
of Posidonia sp. leaves, roots and fibrous leaf sheaths. Organic matter was dried at 60ºC
for 3 days and ground to <2mm using a coffee grinder. Sediments were homogenously
mixed to a known mass of organic matter according to treatment (Table 1). A total of 144
replicate seedlings were each randomly planted in individual pots with a randomly
assigned treatment. Seedlings were cleaned by hand every 2-3 days to ensure no algal
overgrowth that could limit light availability.
47
Fraser (2016)
Post-harvest measurements
All seedlings were harvested after six months. Prior to harvesting (7 days), five replicate
shoots from each treatment within each tank were marked by punching a hole with a pin
through all leaves at the top of each shoot sheath (Short and Duarte 2001). After
harvesting, leaves were removed then separated into old and young leaves with ages based
on position on shoot (oldest leaf with the lowest vertical position on a shoot, and
successively younger leaves positioned in ascending order along the shoot). New leaf
growth was defined as the distance between the initial marking scar and base of the leaf
(young and old leaves), plus any new unmarked leaves (young) formed during the seven
day period. New leaf growth was dried (see above) and weighed for leaf productivity
estimates. Seagrass biomass was determined from five replicates in each tank. Seedlings
were cleaned of epiphytes and separated in component parts (individual leaves, roots, and
rhizomes). These components were then dried at 60ºC for 72 hours and weighed to obtain
biomass. Due to low biomass, leaves from each treatment in each tank were pooled before
nutrient and stable isotope analyses. Leaves were ground to a fine powder using a steel
ball-mill. Carbon (C) and N concentrations were determined using an Automated C/N
Analyser-Mass Spectrometer consisting of a 20/22 mass spectrometer connected to an
ANCA-S1 preparation system (Sercon, Crewe, UK) at the Western Australian
Biogeochemistry Centre at the University of Western Australia. All samples were
standardized using multi-point normalization against a secondary reference of Radish
collegate (3.167% N, 41.51% C,), which was in turn standardized against primary
analytical standards (International Atomic Energy Agency, Vienna) (Paul et al. 2007).
The external error of analyses (one standard deviation) was no more than 0.1 for C:N
ratio. Phosphorus contents of seagrass and sediment samples were determined using a
dry-oxidation and acid hydrolysis technique, followed by a colourimetric analysis of the
phosphate concentration of the extract (Fraser et al. 2012). Phosphate concentrations were
compared with Australian Soil and Plant Analysis Council Standards 102 (Pinus radiata)
and 105 (Eucalyptus sp.), which were treated accordingly. We found this technique to
yield 94-101% of the reported P content of ASPAC Standards 102 and 105. Elemental
concentrations of seagrass and sediment samples were calculated as a percentage of dry
weight, and converted to elemental ratios.
48
Phosphorus cycling in seagrass sediments
Sediment biogeochemistry
The activities of four hydrolytic enzymes were measured using standard colourimetric
methods (Dick 2011). The enzymes and substrates used were (i) α-glucosidase assayed
with 4-Nitrophenyl α-D-glucopyranoside, (ii) β-glucosidase assayed with 4-Nitrophenyl
β-glucopyranoside, (iii) acid phosphatase and (iv) alkaline phosphatase both assayed with
4-nitrophenyl phosphate disodium hexahydrate. Briefly, 1.00 g of sediment was added to
a 50 ml Erlenmeyer flasks with 4 mL of modified universal buffer (pH=6.00 for αglucosidase and β-glucosidase, pH=6.50 for acid phosphatase and pH=11.00 for alkaline
phosphatase) and 1 mL substrate, and were incubated at 37 ºC for one hour. Following
incubation, 1 mL of 0.5M CaCl2 and 4 ml of 0.5M NaOH (for phosphatases) or 4 ml
10mM tris(hydroxymethyl) aminomethane (pH=12, for glucosidases) were added to stop
the reaction. The final nitrophenol concentrations were determined photometrically at 400
nm. Resin extractable P was measured using anion exchange membranes that are a good
representation of P that is available for uptake by plants (Cooperband and Logan 1994;
Cheesman et al. 2010; Bünemann et al. 2013). Membranes were cut into 4 cm x 1 cm
strips, and four were inserted into sediments in each pot. The membranes were left for
four weeks to reach equilibrium with porewaters, before being removed, rinsed in DI
water, and placed in 30 ml 0.5M HCl. Resin extractable P was then measured from HCl
extracts using the Murphy Riley method for colourimetric determination of P
concentrations (Murphy and Riley 1962). Sediment redox potential was measured six
months after sediments were installed using a platinum electrode connected to a pH
electrode and measured as millivolts. Platinum electrodes were left in sediments for 48
hours to allow biogeochemical profiles to be re-established after initial disturbance. No
data was collected for redox potential in 2.5% OM treatments due to logistical difficulties.
Sediment penetration resistance was measured using a Pocket Penetrometer (Humboldt
Mfg. Co., Illinois, US).
Microbial molecular community identification
Surface sediment samples (2 ml, 0-2 cm depth) for DNA extraction and genetic analysis
were collected from pots using a 3 ml syringe. Samples were transferred to a 2 ml
cryovial, and immediately frozen in liquid nitrogen (N) before storage at -80ºC. Sediment
samples were then homogenized using a sterile metal spatula, and 0.5 g was added to a 2
ml Eppendorf tube. DNA was extracted from sediments using the MP Bio Fast DNA
49
Fraser (2016)
SPIN kits for soil (MP Bio, California, United States of America) with some
modifications (Abell et al. 2011). DNA was eluted in nuclease-free water and DNA
extraction was confirmed by agarose gel electrophoresis.
DNA was amplified using universal bacterial primers 16S-1392F (5’GYACACACCGCCCGT-3’) and a 5’HEX labelled version of the reverse primer 23S125R (5’-GGGTTBCCCCATTCRG-3’) (Hewson and Fuhrman 2004). 50 μL reactions
containing 1 μL DNA, 25 μL of GoTaq Green Master Mix (Promega, Australia), 2 μL of
Bovine Serum Albumin (Promega), 1 nM of each primer were amplified as follows: initial
denaturation at 94°C for 3 minutes, followed by 35 cycles of 94°C for 1 min, 55°C for 1
min 30 sec and 72°C for 1 min 30 sec, and a final extension step of 72°C for 10 min.
Successful amplification was confirmed by gel electrophoresis, and PCR products were
purified using AMPure magnetic beads (Agencourt, Beckman Coulter Life Science,
USA) according to the manufacturer’s instructions, resuspended in 40 μL of TE buffer
and quantified using a Nanodrop 8000 spectrophotometer (ThermoScientific,
Wilmington, USA).
Microbial community DNA was analysed by Automated Ribosomal Intergenic
Spacer Analysis (ARISA). Approximately 25-40 ng of purified amplified PCR reaction
was mixed with 0.4 μL of internal size standard (GS1200-Liz, Applied Biosystems) in
deionised formamide and denatured at 94°C for 2 minutes. Denaturing capillary
electrophoresis (on a 50cm capillary array) was carried out using an ABI 3130XL genetic
analyzer (Applied Biosystems). ARISA profiles were analysed using GeneMapper v3.7
(Applied Biosystems), with peak height as a measure of community presence, and a peak
threshold of 50 relative fluorescence units greater than the baseline.
While the drying of sediments likely altered the initial microbial community
composition associated with the sediments, this was necessary to ensure the quantitative
accuracy of each OM treatment. All OM treatments were received the same protocol,
removing any potential treatment effect. Any introduced sampling artefacts would be
minimized by collecting samples for microbial community identification at the end of the
experiment (i.e. six months after being submerged), giving microbial communities time
to restablish.
Statistical analysis
Relationships between seagrass biomass, productivity, nutrient content, sediment enzyme
50
Phosphorus cycling in seagrass sediments
expression and organic matter content were analysed using linear regression analysis.
Prior to analysis, data were checked for normality and homogeneity of variances. The
relationship between the sediment microbial community and measured environmental
variables was investigated using a distance-based multivariate multiple regression model
(distance linear based model [DistLM]) marginal tests with a Bray-Curtis distance matrix
based on 9,999 permutations. Seagrass and sediment biogeochemistry data were analysed
using R version 3.1.1 (R Core Team 2014) using the base program and the ‘ggplot2’
package for graphics (Wickham 2009). Variation in microbial community structure
between samples was analysed in PRIMER v6.0 (Clarke and Gorley 2006) with the
PERMANOVA+ Add-on (Anderson et al. 2008).
Results
Sediment biogeochemistry
Redox potential in sediments strongly decreased with the addition of organic matter (Fig.
3.1a, R2 = 0.86, p<0.0001, n=15), indicating more reducing conditions in sediments.
Redox potential ranged from a maximum of -92.9 mV (± 9.9 mV) in sediments with 0%
OM to -375.4 mV (±14.2 mV) in sediments with 2.0% OM. Penetration resistance in
sediments was significantly greater with the addition of organic matter (Fig. 3.1b, R2 =
0.93, p < 0.0001, n=18), indicating a higher force needed for roots to penetrate into
sediments. Resin extractable P tended to decrease in sediments as the concentration of
OM increased, although this was not statistically significant at p = 0.05 (R2 = 0.15, p =
0.059, n=24, data not shown).
Enzymes involved in C cycling showed variable responses to OM enrichment
(Fig. 3.2). α-glucosidase expression was not significantly related to concentration of OM
in sediments (Fig. 3.2a, R2=0.08, p=0.13, n=18). β-glucosidase expression significantly
increased as the concentration of OM in sediments increased, indicating the potential for
increased cellulose degradation (Fig. 3.2b, R2 = 0.79, p<0.0001, n=24).
51
Fraser (2016)
Fig. 3.1 Relationship between sediment organic matter % and (a) sediment redox potential (n=3) and (b)
sediment penetration resistance (n=3). Black dots represent data points, dashed lines represent linear models
fitted to the data, and the grey outline shows 95% confidence intervals. Note that no data was collected for
redox potential in 2.5% OM treatments
Enzymes involved in releasing phosphate from OM had higher expression in
sediments with OM added, in comparison to the 0% OM treatment. Alkaline phosphatase
expression also significantly increased as OM content in sediments increased (Fig. 3.2c,
R2=0.45, p=0.0009, n=24). Acid phosphatase expression significantly increased as
sediment OM concentration increased (Fig. 2d, R2=0.53, p<0.0001, n=24).
52
Phosphorus cycling in seagrass sediments
Fig. 3.2 Expression of the enzymes α-glucosidase (a, n=3), β-glucosidase (b, n=4), alkaline phosphatase (c,
n=4), and acid phosphatase (d, n=3) in sediments with Posidonia australis seedlings ranging from 0% to
2.5% OM (d.w.). Enzyme activities are in µmol-1 substrate consumed hr-1 g-1 dry sediment. Black dots
represent data points, dashed lines represent linear models (where significant) fitted to the data, and the
grey outline shows 95% confidence intervals
Microbial community structure
The sediment microbial community composition was significantly altered upon
enrichment with OM (PERMANOVA p<0.001). Pairwise differences between treatments
tended towards significance where differences in OM content were largest. For example,
there were no significant pairwise differences between the 0%, 0.5%, 1%, and 1.5% OM
treatments, whereas the 2.5% OM treatment was significantly different to all treatments
53
Fraser (2016)
except 1.5% OM, which showed overlap with communities from enriched and nonenriched treatments.
The remaining environmental data measured in the experiment (hydrolytic enzymes,
sediment redox potential, and resin P) only accounted for 16% of the total variation in the
microbial community (Fig. 3.3). Changes in microbial communities were most strongly
correlated to changes in sedimentary organic matter content (accounting for 12.2% of the
variation), β-glucosidase expression (10.8% of total variation), and sediment redox
potential (10.6% of total variation). All other variables accounted for <2% of the variation
in the microbial community (i.e. over 50% of the microbial community variability is
unexplained by our measured variables).
Fig. 3.3 Relationship between environmental data and changes in sediment microbial communities
measured by ARISA. Numbers and shading denote organic matter treatment (0 – 2.5% OM). Plot represents
a distance based redundancy analysis (dbRDA) ordination of ARISA trace data derived from Bray-Curtis
distance matrix and with environmental variables chosen by stepwise selection. The axes plotted account
for 16.4% of the total variation
Seagrass biomass
Total seagrass seedling biomass did not significantly change with OM amendments (Fig.
3.4a, R2 = 0.003, p = 0.092, n=90). Similarly, seedling leaf biomass did not change
significantly with OM addition (Fig. 4b, R2 = 0.01, p = 0.302, n=90). Organic matter
54
Phosphorus cycling in seagrass sediments
enrichment did not change total leaf productivity (Fig. 3.4c, R2 = 0.01, p = 0.47, n=90) or
productivity in young leaves (R2 < 0.01, p = 0.9, n=90, n=19). There were strong
relationships in the biomass allocation between different leaves among OM treatments.
Biomass was significantly greater in older leaves at lower OM concentrations (Fig. 3.4d,
R2 = 0.32, p < 0.0001, n=90), while biomass was significantly greater in newer leaves at
higher OM concentrations (R2 = 0.13, p = 0.016, n=90).
Fig. 3.4 Total seedling biomass (a), total leaf biomass (b), leaf productivity (c), and leaf specific biomass
(d) for Posidonia australis seedlings growing in sediments with organic matter concentrations ranging from
0% to 2.5% OM (d.w.). Values are means from 15 replicates, with bars showing ± 1 s.e.
Rhizome biomass decreased significantly with sediment OM enrichment, although
this relationship was weak (Fig. 3.5a, R2 = 0.059, p = 0.012, n=90) and non-linear. Root
biomass also decreased significantly with OM enrichment; again this relationship was
55
Fraser (2016)
weak and non-linear (Fig. 3.5b, R2 = 0.047, p = 0.023, n=90). Total belowground biomass
decreased significantly upon OM addition (Fig. 3.5c, R2 = 0.078, p = 0.008, n=90). The
decrease of root and rhizome biomass at high OM led to a significant increase in
above:belowground biomass with OM enrichment (Fig. 3.5d, R2 = 0.046, p = 0.024,
n=90).
Fig. 3.5 Total rhizome biomass (a), total root biomass (b), total belowground biomass (c), and aboveground:
belowground biomass ratios (d) for Posidonia australis seedlings growing in sediments with organic matter
concentrations ranging from 0% to 2.5% OM (d.w.). Values are means from 15 replicates, with bars
showing ± 1 s.e.
Seagrass nutrient concentrations
The nutrient concentration of seagrass leaves was variable with OM enrichment (Table
56
Phosphorus cycling in seagrass sediments
3.1). Carbon (R2 = 0.05, p = 0.78, n=18), nitrogen (R2 = 0.15, p = 0.06, n=18), and
phosphorus (R2 = 0.13, p = 0.08, n=18) concentrations in seagrass leaves did not
significantly change with OM addition. However, the carbon:nitrogen (C:N) ratio of
seagrass leaves significantly increased with increasing OM concentrations (R2 = 0.41, p
= 0.0042, n=18). Carbon:phosphorus (C:P) ratios of seagrass leaves tended to increase
with elevated OM concentrations, though this relationship was weaker and not significant
at the 0.05 level (R2 = 0.20, p = 0.062, n=18). Nitrogen:phosphorus (N:P) ratios of
seagrass leaves showed no significant relationship to OM concentrations (R 2 = 0.02, p =
0.2768, n=18). No significant relationships were detected between sediment OM and total
seagrass nutrient content after correction for seedling biomass (i.e. total leaf biomass x
C/N/P (%d.w.)).
Table 3.1 Carbon (C), nitrogen (N), and phosphorus content (% d.w), and carbon:nitrogen (C:N),
carbon:phosphorus (C:P), and nitrogen:phosphorus (N:P) ratios of seagrass leaves growing in sediments
with different concentrations of organic matter. Values are means  standard error (in parentheses).
Sedimentary organic matter concentration (% dry weight)
0%
0.5%
1%
1.5%
2%
2.5%
C (% d.w)
37.3 (0.3)
37.5 (0.1)
37.1 (0.3)
37.0 (0.2)
37.3 (0.5)
37.3 (0.2)
N (% d.w)
2.0 (0.1)
2.0 (0)
1.9 (0.1)
1.9 (0.1)
1.9 (0.1)
1.9 (0.1)
P (% d.w.)
0.15 (0.01) 0.16 (0.01) 0.14 (0.01) 0.15 (0.002) 0.14 (0.004) 0.13 (0.02)
C:N
18.6 (0.3)
18.6 (0.2)
19.4 (0.7)
19.4 (0.5)
19.8 (0.1)
20.0 (0.4)
C:P
248 (17)
226 (16)
269 (12)
239 (4)
268 (7)
287 (33)
N:P
13.4 (1.1)
12.1 (0.9)
13.9 (0.1)
12.4 (0.4)
13.6 (0.30) 14.3 (1.4)
Discussion
Seagrass detritus is a ubiquitous component of OM in seagrass sediments.
Consequently, established seagrass meadows have been considered as hotspots for C
burial over long temporal and spatial scales (Fourqurean et al. 2012), but the ecological
importance of this OM over smaller temporal and spatial scales is largely unknown. Here,
we show the significant impacts of seagrass-derived OM on the biogeochemistry and
microbial communities within sediments, and on biomass partitioning in P. australis
seedlings.
57
Fraser (2016)
Elevated seagrass OM concentrations in sediments strongly increased the
expression of hydrolysing enzymes in sediments, confirming our first hypothesis. The
enzymes examined were those involved with C and P release from complex organic
molecules (Fig. 3.2). We found that the addition of OM led to distinct changes in
biogeochemical conditions in sediments between treatments, which could alter sediment
nutrient availability. Resin P tended to decrease with OM enrichment, and increased
production of phosphatase indicated a likely alteration in the physiological demands of
the microbial community. For example, the addition of OM could have released the
microbial community from C limitation, shifting demands onto other nutrients (e.g. P)
(Haugwitz and Michelsen 2011). The immobilization of P and N into microbial biomass
could then reduce the concentrations in porewaters, resulting in increased plant-microbial
competition for nutrients. An increase in plant-microbe competition for nutrients after
OM enrichment has been well established in agricultural systems (Lehmann et al. 2011;
Quilliam et al. 2012). Alternatively, the greater aboveground biomass in OM enriched
treatments may have led to a dilution of potential nutrient impacts, though the necessary
pooling of samples for nutrient analysis prevent us from exploring in further detail.
Regardless, the ecological role of seagrass detritus on sediments is poorly known, and has
largely been considered as a refractory pool of C due to low turnover rates (Mcleod et al.
2011). We suggest that in natural meadows, the deposition of detritus may increase
competition for nutrients between seagrasses and microbes as meadows age and deposit
more OM into sediments, shifting nutrient limitations as meadows develop. However, the
extent of nutrient limitation would depend on other nutrient inputs, such as sedimentation
of suspended materials, which are also enhanced by seagrass cover (Risgaard-Petersen
and Dalsgaard 1998; Gacia and Duarte 2001). Clearly, the ecological role that seagrass
detritus plays in coastal sediments must be considered when examining the
biogeochemistry of seagrass meadows.
There were significant shifts in the sedimentary microbial community as OM
concentrations increased, confirming our second hypothesis. Increased seagrass OM
concentrations may have benefitted microbial groups that were actively able to utilise the
substrate by releasing enzymes (Coolen et al. 2011; Khodadad et al. 2011), or by altering
environmental conditions (Allison and Martiny 2008). We observed here that seagrass
58
Phosphorus cycling in seagrass sediments
OM addition resulted in more reduced sediments (Fig. 3.1a), which would favour
microbes that require little or no oxygen, such as sulphate reducers (Rosselló-Mora et al.
1999). These bacterial taxa play an important role in OM remineralization in marine
sediments, and within the plant rhizosphere may even form mutualistic relationships with
seagrasses, due to the capacity of many sulphate reducers to fix N, thereby making it
available for uptake by the plant (Welsh 2000). It appears that 1.5% OM in sediments
may represent a threshold concentration, after which a significant shift in the microbial
community composition occurs. It is unknown whether such a threshold would also hold
true in situ; especially given the capacity of large seagrass root systems to oxygenate
sediments (Greve et al. 2003; Pedersen et al. 2004; Borum et al. 2005), therefore buffering
environmental conditions and, presumably, microbial metabolic processes. Given that
sulphide intrusion from sediments has been suggested as a contributor of seagrass decline
(Borum et al. 2013; Holmer and Hasler-Sheetal 2014), the role of sulphate reducers in
seagrass sediments warrants further investigation.
There were modest decreases in root and rhizome biomass in seedlings growing in
enriched OM treatments (Fig. 3.5). Posidonia australis exhibits strong architectural and
morphological plasticity in root growth under changing environmental conditions. For
example, P. australis growing in summer at Oyster Harbour, Australia had more lateral
roots than in winter (Hovey et al. 2011; Hovey et al. 2012). Drivers of root plasticity are
well known in many terrestrial plants, but factors that influence root growth and
morphology in seagrasses are poorly characterised, despite the importance of seagrass
roots for nutrient uptake and anchorage (Alcoverro et al. 2000; Belzunce et al. 2008). Our
study shows that P. australis seedlings demonstrate plasticity in root growth even at an
early developmental stage (<6 months after settling) when exposed to variable OM
concentrations, even if the mechanisms for this are still unknown. Signalling compounds
such as hormones may play a role in seedling root development, though have shown
variable results in previous experiments with Posidonia spp. (Balestri and Bertini 2003;
Glasby et al. 2015).
Seagrass OM addition led to limited changes in aboveground seedling biomass and
productivity (Fig. 3.4, Table 3.1). The use of nutritional reserves in seedlings or uptake
of nutrients from seawater may have homogenised responses to increased OM, even
59
Fraser (2016)
where some treatments may confer an ecological advantage over a longer timescale.
Though productivity did not significantly change between treatments, seedlings allocated
more biomass to old leaves when growing in low OM sediments, and to new leaves when
growing in high OM sediments. This suggests changes in leaf senescence rates. Leaf
senescence is a plastic trait that is highly dependent on environmental conditions such as
nutrient availability (Guo and Gan 2005). By signalling a high availability of C relative
to N in old leaves, sugar accumulation triggers leaf senescence, remobilizing N into new
leaves and preventing loss of N from the plant (Wingler 2005). Nitrogen accumulation
was lower in high OM treatments (Table 3.1), suggesting that a decrease in N availability
driven by OM additions may alter leaf senescence rates.
We did not find significant evidence supporting the previous hypotheses that
seagrass derived OM assists P. australis seedlings through sediment microbial activity
(Statton et al. 2013), but there are other ecological advantages that higher seagrass OM
concentrations in sediments may provide to seedlings. Though sediment penetration
resistance was greatest in sediments with higher OM concentrations (Fig. 3.1b), the level
of physical force required to remove seedlings from sediments was also greater in
sediments with higher OM concentrations (J. Statton, pers. comm). Sediment OM
enrichment increases secondary and tertiary branching of P. australis roots (Statton et al.
2013), which increase stability in sediments (Grossnickle 2005). As a result, the presence
of OM may assist seedlings in early survival by limiting hydraulic removal; a major
bottleneck to seedling survival (Kirkman 1999; Statton et al. 2012). An increase in
anchorage potential in organic rich sediments would have implications for seagrass
recovery following dieback events that can input large volumes of seagrass-derived OM
into sediments (Walker et al. 2006; Fraser et al. 2014). This may also have restoration
implications, as hydrodynamic scouring is a major problem in seagrass restoration
projects (Statton et al. 2012). Choosing sites with higher seagrass OM in sediment or
preconditioning sediments with seagrass OM prior to transplanting may provide means
of increasing restoration success.
The addition of seagrass OM to sediments also changed other physical conditions
that could have influenced seedling health. Sediments with higher OM concentrations had
significantly more negative redox potentials (Fig. 3.1a), indicative of lower oxygen
60
Phosphorus cycling in seagrass sediments
concentrations resulting from increased respiration rates. Increased OM concentrations in
seagrass sediments also increase sulphide in porewater, which can contribute to seagrass
dieback through sulphide intrusion (Borum et al. 2005). Seagrasses are particularly
susceptible to sulphide intrusion during times of low oxygen concentrations. However,
we found no decrease in seagrass health or survival in high OM treatments in this
experiment. The contribution of sedimentary organic matter to toxic sulphide intrusion
may be overestimated due to the nature and concentrations of OM added in previous
experiments. For example, experimental additions of OM are often in unique forms that
would not be found in significant concentrations in natural seagrass sediments such as
sugar (Pérez et al. 2007) or sodium acetate (Ruiz-Halpern et al. 2008), and therefore
would not be ecologically representative of in situ conditions. Similarly, conclusions
drawn from natural experiments have often been conducted in highly degraded systems
such as fish farms with anthropogenically derived OM additions that again may not
represent conditions that are typically found in healthy seagrass meadows (Delgado et al.
1999). The dominant source of OM in most natural seagrass meadows would be seagrass
detritus, especially in meadows dominated by large, long-lived species such as P.
australis (Fourqurean et al. 2012). Seagrass detritus is notoriously refractory in nature.
The high difference in lability between the OM additions in this and other experiments
would result in very different below ground impacts - with differences in decomposition
rates, and microbial and environmental shifts resulting from the additions. This is
evidenced by the lack of negative impacts on seedlings here and in similar experiments
(Statton et al. 2013), and the often-abundant growth of seagrasses upon sediments
naturally enriched with seagrass derived OM. For example, seagrasses form highly
productive meadows in blue carbon hotspots like the Mediterranean Sea and Shark Bay,
Western Australia, even where sedimentary sulphide concentrations are high (Cambridge
et al. 2012). The ecological consequences of OM enrichment of seagrass sediments will
be highly dependent on the types and quantities of OM added. Future experiments should
focus on adding OM types and concentrations that are representative of natural conditions
if ecologically relevant conclusions are desired.
61
Fraser (2016)
Acknowledgements
We thank the support staff at the Department of Fisheries, WA for logistical support
throughout the experiment. We also thank the CSIRO Marine and Atmospheric Research
(CMAR) Environmental Genomics team in Hobart, Tasmania, for their help with
microbial community analyses; in particular Dr Guy Abell. This project was funded
through Australian Research Council Linkage Grants to GAK and JS (LP100200429 &
LP130100155). BL was funded by an Indian Ocean Marine Research Centre (IOMRC)
Postdoctoral Fellowship, supported by AIMS, CSIRO and the University of Western
Australia (UWA). MF supported by Ph.D. funding from the School of Plant Biology,
UWA. We thank 3 anonymous reviewers who provided thoughtful comments that greatly
improved the manuscript.
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Chapter 4: Organic phosphorus accumulates in calcareous
seagrass sediments with increasing salinity
Matthew W. Fraser, Gary A. Kendrick, Grzegorz D. Skrzypek, Lindsay T. Byrne,
Pauline F. Grierson
Preamble: This manuscript will be submitted to Limnology and Oceanography, and is
formatted to the requirements of the journal. The work is primarily my own. GAK
received funding for the project. GAK and PFG assisted with experimental design. LB
provided help with NMR analysis and interpretation. GAK, GS, and PFG all provided
feedback on the manuscript. This chapter examines the different fractions of sedimentary
phosphorus across the salinity gradient in Shark Bay sediments, including examining the
organic P pool to the compound level using 31P-NMR, providing a unique insight into the
organic P pool in seagrass ecosystems.
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Phosphorus cycling in seagrass sediments
Abstract
Seagrasses growing on calcareous sediments in oligotrophic waters are often assumed to
be phosphorus (P) limited given the high P adsorption capacity of these sediments.
However, little is known of P speciation in seagrass sediments and if seagrasses are able
to utilize P from other fractions such as calcareous bound-P or reducible P complexes in
sediments. In particular, mineralization of organic P from seagrass litter could play an
important role in reducing P limitation in oligotrophic seagrass meadows. Here, we
sought to quantify the amounts and proportion of inorganic and organic fractions in
calcareous sediments of Shark Bay, a subtropical embayment characterised by large
seagrass meadows and a strong salinity gradient. We used a sequential extraction method
to examine P fractions in sediments across a range of salinities (39-47‰). We also used
31
P-nuclear magnetic resonance (NMR) to characterise the sediment organic P fraction to
the compound level - the first time this has been done is seagrass sediments. We found
that the most available P fractions (magnesium extractable P and reducible P) were not
significantly related to salinity, but tended to be greater at low salinity sites, while the
proportion of organic P in sediments significantly increased with salinity. The proportion
of organic P was highly correlated with sedimentary organic matter content (R2 = 0.50,
P < 0.0001). The only organic P compounds detected by NMR were phosphate
monoesters, which are a readably mineralisable form of P and thus a potential source of
P to seagrass. We propose that cycling of seagrass detritus is of increasing importance in
maintaining P requirements of seagrasses as salinity increases and soluble reactive
phosphorus in the water column decreases.
Keywords: 31P-nuclear magnetic resonance, nutrient availability, phosphate monoesters,
sediment organic phosphorus, SEDEX, Shark Bay
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Fraser (2016)
Introduction
Seagrasses growing in oligotrophic coastal habitats with calcareous sediments and low
phosphorus (P) inputs are often considered P limited because of low soluble reactive P
(SRP) concentrations in overlying waters (Short 1987). However, root uptake of P from
sediments can supply a substantial proportion (up to 100%) of the P requirements of
seagrasses, particularly when growing in ecosystems with low concentrations of SRP in
the water column (Jensen et al. 1998; Gras et al. 2003). Phosphates in seagrass sediments
are present as a diverse range of compounds that differ in their availability for biological
uptake (Koch et al. 2001; Holmer et al. 2006), yet for the most part are undescribed. The
organic P fraction in particular is poorly understood in marine sediments, despite growing
recognition of its ecological importance and complexity in terrestrial soils and freshwater
sediments (Turner et al. 2005; Baldwin 2013). In seagrass sediments of Florida Bay,
organic P can comprise almost half of the total sediment P (Koch et al. 2001). However,
the chemical composition of organic P and therefore its potential mineralization and
availability for uptake by plants is unknown. Further, the environmental drivers
influencing the spatial distribution of different P fractions across marine ecosystems,
including coastal embayments, are not well known.
Previous studies have shown that sediment P in coastal ecosystems can be strongly
influenced by salinity of the overlying water column. Shifts in P fractions of seagrass
sediments have been examined along freshwater-marine transitions to salinities from 537‰ at Florida Bay (Koch et al. 2001). This study found that porewater P did not increase
with salinity, given the high adsorption capacity of the predominately calcareous
sediments. Elsewhere, in an analysis of a temperate estuary in South Carolina, porewater
P increased as salinity increased, but sediment sorbed P (measured as bicarbonateextractable P) decreased (Sundareshwar and Morris 1999). However, this study did not
assess concurrent change in organic P. Other studies have found that under high salinity,
organic P compounds such as phosphate monoesters may become more mineralizable due
to increased desorption from sediment surfaces, and thus available for root uptake
(Monbet et al. 2009). However, it remains unclear if and how the distribution of P forms
in calcareous sediments is affected under hypersaline (>40‰) conditions and how this
may influence benthic production, limiting our understanding of P cycling and nutrient
limitation in these ecosystems.
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Phosphorus cycling in seagrass sediments
Different fractions of sedimentary P can be operationally grouped based on sequential
extractions, which provide a more comprehensive assessment of overall P availability
than simply measuring single extracts of inorganic P or total P. One such extraction
technique is SEDEX (Ruttenberg 1992), which has been developed and modified for
specific use with marine sediments. The SEDEX technique splits total sedimentary P into
four operationally defined pools that differ in their bioavailability; loosely sorbed P,
reducible P, calcium-bound P, and residual organic P. While sequential extraction
methods necessarily define P pools based on the types of extractant used, the results can
provide ecologically relevant information. For example, loosely sorbed P primarily
consists of inorganic phosphate that is readily available for uptake by plants, whereas
calcium-bound P is relatively unavailable for direct uptake by plants (Short 1987).
Reducible P comprises P adsorbed onto reducible metal oxyhydroxides, and is strongly
affected by the concentrations of metals in sediments such as Fe, Mg and Al, as well as
sulfur that can sequester Fe and potentially release P (Koch et al. 2001). Splitting
sedimentary P into such fractions provides detailed information about the potential
availability of sedimentary P for biological uptake, therefore allowing more accurate
predictions for biological responses to external P inputs.
Shark Bay is a large sub-tropical marine embayment with extensive seagrass communities
that form some of the largest continuous meadows in the world (Kendrick et al. 2012).
Salinity across Shark Bay ranges from 35‰ in the north to ~70‰ in the south (Kendrick
et al. 2012). While the relationship between seagrass physiology and growth is reasonably
well understood for Shark Bay (Walker 1985; Kendrick et al. 2012), remarkably little is
understood of the interactions between seagrass distributions and sediment
characteristics. However, sediments are known to be predominantly calcareous (mean =
71.3% CaCO3) and sediment organic matter content (OM) also increases with salinity
(Fourqurean et al. 2012). SRP concentrations in the water column are low throughout
Shark Bay and frequently below detectable limits (<0.02 μM) in its southern, hypersaline
reaches (Atkinson 1987; Fraser et al. 2012). As such, seagrasses in Shark Bay are
considered to be P limited, especially in hypersaline areas (Burkholder et al. 2013).
However, sediments may also represent a source of P to seagrasses given the dominant
species (Amphibolis antarctica and Posidonia australis) can absorb nutrients via their
root systems (Pedersen et al. 1997; Alcoverro et al. 2000). The highly variable P content
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Fraser (2016)
of seagrass leaves observed over small spatial scales in Shark Bay (~10km; Fraser et al.
2012) suggests that (i) root uptake from sediments at least partially meets P demand of
these plants and (ii) that the distribution of labile P fractions in the sediments is
heterogeneous.
In this study, our objective was to determine how the total and proportional concentrations
of different phosphorus fractions varied in sediments along the salinity gradient in Shark
Bay, and evaluate the potential of sediment as a P source for seagrasses. We used
sequential extraction and 31P-NMR to determine the concentrations of different sediment
P fractions and to identify dominant organic P compounds. Specifically, we hypothesized
that: (i) calcareous-bound P will be the dominant fraction throughout the majority of
sediments in Shark Bay, given that calcareous sediments dominate; (ii) concentrations of
labile P (magnesium-extractable P and reducible P) would be lower and comprise
proportionally less of the total sediment P at hypersaline sites, given the lower overlying
SRP in the water column; (iii) concentrations of organic P fractions would be highest and
comprise a greater proportion of total sediment P at high salinity sites, given that sediment
OM increases with salinity (Fourqurean et al. 2012). We also identified organic P
compounds across the salinity gradient in Shark Bay using 31P-NMR .
Methods
Site description & experimental design
Eight sites were selected across the eastern embayment of Shark Bay (Fig. 4.1). All sites
~1.5 km apart and had overlying water of 1-2m deep. Three sites (sites 6-8) were adjacent
to the mouth of the ephemeral Wooramel River, the only river that flows directly into
Shark Bay. The benthic community at all sites was dominated by Amphibolis antarctica
with Halodule uninervis growing as an understory species. Salinity of the water column
was determined at each site with a TPS 90FL-T analyser (TPS Pty Ltd, Springwood,
Queensland, Australia). Salinity ranged from 37‰ at site 1 to a maximum of 46.7‰ at
site 7 (Table 1).
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Phosphorus cycling in seagrass sediments
<40 psu
40 - 44 psu
>44 psu
N
Figure 4.1: Location of the study site at Shark Bay, Western Australia. Black circles denote sites.
Haloclines showing salinity at time of sampling are shown in dashed lines, with corresponding salinity
ranges measured in this study in italics. Inset shows location of Shark Bay.
Sediment cores
At each site, six replicate sediment cores (50 mm diameter x 500 mm length) were
collected in March 2013. PVC Cores were driven into the sediment using a hammer, then
carefully removed from the sediment, capped at both ends, and placed in a cooler with
ice. Cores were frozen prior to transport back to the laboratory in Perth (800 km away) to
prevent multiple thawing and freezing of samples that can influence P concentrations
(Vaz et al. 1994). We acknowledge that this treatment may result in overestimation of
inorganic phosphate and underestimation of pyrophosphates (Turner et al. 2007), but was
unavoidable given the remoteness of the site. Each core was divided into six different
depth sections: 0-2 cm, 2-5 cm, 5-10 cm, 10-20cm, 20-30 cm, and 30-40 cm; each size
section was then homogenized by mixing the sample in a sealed bag. Large seagrass
73
Fraser (2016)
fragments and shells (>5mm) were removed. A 1 g sub-sample was taken from each bag
for sequential P extraction, while the remainder was used for chemical and physical
characterisation of sediments.
Sediment chemical and physical characterization
Sediment grain size fractions were determined using a dry sieving technique using 1180,
850, 500, 355, 250, 106 and 63 μm sieves (Endecotts Ltd.). The OM content of sediments
was estimated from loss on ignition (LOI), which was determined by the difference in
sediment weight after combustion of 1g sediment samples at 550 °C for 8 hours
(Fourqurean et al. 2012). Elemental analyses of sediment NaOH/EDTA extracts (Al, Ca,
Fe, Mg, Mn, P, S, and As) were performed on a PerkinElmer Optima 5300DV inductively
couple plasma optical emission spectrometer (ICP-OES).
Sequential P extraction of sediments
We used a sequential extraction method for marine sediments (Ruttenberg 1992), with
modifications for carbonate-dominated sediments developed by Jensen et al. (1998) and
Koch et al. (2001). Magnesium extractable P, which is considered the most biologically
available fraction, was extracted using 49 mL of 1 M MgCl2 (pH = 8.00). 1 mL of 2 M
sulfuric acid was added to the sample extracts for storage prior to P measurement. 50 mL
of bicarbonate-dithionite solution (0.11 M NaHCO3 and 0.11 M Na2S2O4, pH = 7.00) was
then added to the residues from the previous procedure to extract reducible P adsorbed
by crystalline and amorphous iron and manganese compounds (Reducible P) (Jensen and
Thamdrup 1993). Prior to preservation, the supernatant was bubbled with nitrogen gas
(N2) for 20 minutes and then the supernatant preserved with the addition of 4 mL of 1 N
sulfuric acid, and 21 mL of deionised water. Preservation resulted in the precipitation of
elemental sulfur, necessary to exclude sulfides and limit interference during analysis
(Koch et al. 2001). Sediment residues from the bicarbonate-dithionite extraction were
then extracted with 30 mL of 1 N HCl to measure calcareous-bound P (calcareous P).
Sediments were shaken several times and purged to release pressure prior to extraction,
until the release of CO2 stopped. 20 ml of deionised water was then added to the samples.
Residual organic phosphorus was then extracted from the remaining sediment using a
dry-oxidation acid hydrolysis technique (Residual organic P) (Solórzano and Sharp
1980). Residual organic P comprises the majority of sediment organic P, although small
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Phosphorus cycling in seagrass sediments
amounts of labile organic P compounds are likely to be lost in the preceding bicarbonatedithionite extraction (Wang et al. 2013). After each extraction, inorganic phosphate
concentrations were measured spectrophotometrically using a modified ascorbic acid
method (Murphy and Riley 1962; Kuo 1996).
Solution 31P-NMR spectroscopy
Surficial sediments (0-10 cm) were extracted with a standard NaOH/EDTA reagent and
analysed by solution
31
P-NMR spectroscopy following the approach of (Cade-Menun
and Liu 2014). Phosphorus was extracted by adding 20 g of sediment to a 50 mL
centrifuge tube with 40 mL of 0.25M NaOH + 0.05M EDTA solution. The high ratio of
sediment to extractant was necessary to obtain adequate NMR signals. Samples were then
shaken for 16 hours, centrifuged (2000 rpm for 30 minutes) and then filtered through a
Whatman #42 filter, with the supernatant being retained. Samples of the supernatant were
then lyophilized and 1 g of freeze-dried extract was dissolved in 3.6 ml of deionised water,
centrifuged (1400 rpm for 20 minutes), and then added to a 10 mm NMR tube with 0.4mL
of deuterated water (D2O).
31
P-NMR spectra were determined on each sample using a Bruker AV500 (Bruker,
Germany) spectrometer at 298ºK. Data were acquired with a 90º pulse of 41.0 µs,
relaxation delay of 10 s and an acquisition time of 1 s. Typically 5440 scans were acquired
over 16 h 43 min. Chemical shifts were referenced to an orthophosphoric acid standard.
The spectra presented have a line broadening of 5 Hz, and are 1H decoupled.
Data analysis
All data analyses were performed in R version 3.1.1 (R Core Team 2014) using the base
program and ‘ggplot2’ package (Wickham 2009). Relationships between sediment
characteristics, salinity, and sediment P fractions were determined using linear regression
analysis. Relationships between sediment depth, salinity, and P fractions were determined
using two-way ANOVA. Data were log transformed where normality and/or
heterogeneity assumptions were not met. Peak assignment in
31
P-NMR was based on
previous reported procedure (Turner et al. 2003; Cade-Menun et al. 2005; Cade-Menun
and Liu 2014). Changes in the organic P pool at the compound level were analysed
75
Fraser (2016)
visually based on 31P-NMR spectra, as quantitative analysis was not possible given the
low signal:noise ratio encountered in these samples.
Results
General sediment variation along salinity gradient
Shark Bay sediments were predominately composed of particles in the size class 250-500
µm; however finer particles (106-250 µm) were more dominant at sites 1 and 7 (Table
4.1). Organic matter content ranged from 1% in surface sediments at site 1 (mean = 2.3%)
to over 16% at 2-5cm depth at site 6 (mean = 11.0%). Organic matter concentration in
sediments significantly increased with salinity (R2 = 0.63, P <0.0001, n = 144).
Calcium concentrations of sediments ranged between 2193 µg g-1 at site 3 to 3137 µg g-1
at site 1 (Table 4.2). Concentrations of iron and aluminium in sediments were also higher
at higher salinity sites. Magnesium concentrations at site 8 were almost six-fold higher
than any other site. Manganese was only detected at high salinity sites (6-8); sulfur
concentrations were also highest at these sites. Arsenic was highest at sites 7 and 8, and
lowest at site 2. Sediment sulfur concentrations were positively correlated with salinity
(R2 = 0.8, P <0.002, n = 8).
Sediment P fractions
The concentrations and relative proportions of different sediment P fractions changed
significantly across the salinity gradient in Shark Bay (Table 4.2). Both residual organic
P concentration and proportion of total sediment P significantly increased with salinity.
In contrast the concentration and proportion of calcareous P (notionally the least labile
fraction) significantly decreased from >65% (60 μg P g-1) to <40% (23 μg P g-1) along
the salinity gradient. Reducible P concentrations and proportions were weakly positively
correlated to salinity. However, the proportion of magnesium-extractable P – the most
available for biological uptake – did not change in concentration or as a proportion of
total P along the salinity gradient.
Magnesium-extractable P was the smallest P pool in Shark Bay sediments, contributing
an average of 7.4% to the total P in Shark Bay sediments. Magnesium-extractable P
concentration was not significantly related to salinity (R2 = 0.02, P = 0.13, n = 144).
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Phosphorus cycling in seagrass sediments
Similarly, the proportion of total P present as magnesium-extractable P did not differ
across the salinity gradient (R2 = 0.04, P = 0.2, n = 48). Magnesium-extractable P was
highest in sediments at intermediate salinities, with a mean of 8.9 μg P g-1 (9.2% of total
P) at site 4 and 7.7 μg (9.4%) at site 5. While not related to salinity overall, the highest
salinity sites (sites 6, 7 and 8) consistently had the lowest concentrations of magnesiumextractable P (Fig. 4.2). Site 8 had the lowest mean concentration of magnesiumextractable P (2.2 μg P g-1, 4.4%), while site 6 had the lowest concentrations at 20-30 cm
(1.2 μg P g -1; 3.8%) and >30 cm depths (1.2 μg P g-1; 2.4%). Overall, magnesiumextractable P significantly decreased with sediment depth (F5,138 = 47.0, P < 0.0001, Fig.
4.2), but there was no significant interaction between magnesium-extractable P with
salinity and sediment depth.
77
Table 4.1: Summary data for water and sediment chemical and physical characterization at each site. Sediment % fines is the percentage
of sediments with a grain size under 106μm. Sediment aluminium (Al), calcium (Ca), iron (Fe), magnesium (Mg), manganese (Mn), sulfur
(S), and arsenic (As) concentrations are all in µg g-
Fraser (2016)
78
Table 4.2: Mean sediment P fractions and proportions of total sediment P across Shark Bay, averaged across all sediment depths. PMgCl = Magnesium
extractable P. All concentrations are in μg P g DW-1, with standard errors in brackets (N=18).
Phosphorus cycling in seagrass sediments
79
Fraser (2016)
Reducible P pool comprised an average of 11.6% of total P in Shark Bay sediments (Table
4.2, Fig. 4.2). Reducible P ranged from 4.84 μg P g -1 at site 3 to 10.6 μg P g -1 at site 4.
There was a weak positive relationship between salinity and reducible P proportions (R2
= 0.17, P = 0.003, n = 48). Reducible P concentrations were more strongly correlated with
depth than salinity, significantly decreasing as sediment depth increased (F5,138 = 20.45,
P < 0.0001, Fig. 4.2), from a mean of 11.9 μg P g -1 (16.2%) at 0-2 cm to 4.4 μg P g -1
(7.7%) at depths over 30 cm. There was no significant relationship between sediment
reducible P concentrations and Fe (R2 = 0.01, P = 0.75, n = 9) or Al concentrations (R2 =
0.04, P = 0.62, n = 9).
Site 1
Site 2
Site 3
Site 4
Site 5
Site 6
Site 7
Site 8
Depth (cm)
0−2cm
2−5cm
5−10cm
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30−cm
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10−20cm
20−30cm
30−cm
0−2cm
2−5cm
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10−20cm
20−30cm
30−cm
0−2cm
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30−cm
0
Fraction
30
60
90
0
30
Sediment phosphorus (ug P g −1 )
MgCl extractable P
Reducible P
Calcareous P
60
90
Residual organic P
Figure 4.2: Mean sediment phosphorus fractions across Shark Bay, separated by site and depth. White bars
represent magnesium extractable P, light gray bars represent reducible P, dark grey bars show calcareous
bound P, and black bars denote residual organic phosphorus.
80
Phosphorus cycling in seagrass sediments
Residual organic P was the second largest sediment P fraction at all sites, comprising on
average 26.4% of P in sediments. Residual organic P concentrations significantly
increased with salinity (R2 = 0.29, P < 0.0001, n = 144), and the proportion of total P
present as residual organic P also increased with salinity (R2 = 0.66, P < 0.0001, n = 48).
Residual organic P concentrations were highest at intermediate salinities sites, with a
maximum mean of 25.5 μg P g -1 (28%) at site 4 (Table 4.2). However, residual organic
P made up the largest proportion of total P in high salinity sites (sites 6-8, 35-40%), owing
to the lower total P at these sites. Conversely, residual organic P was proportionally
lowest at the low salinity sites (sites 1-3), comprising only 16-17% of total P. There was
no strong relationship between sediment depth and residual organic P concentrations
(F5,138 = 1.22, P = 0.3), which remained relatively stable across all depths (Figure 2). The
concentration of residual organic P in sediments was positively correlated to the OM
content of sediments (R2 = 0.50, P < 0.0001, n = 144, Fig. 4.3a). There was also a strong
positive relationship between the proportion of residual organic P in sediments and sulfur
concentrations (R2 = 0.98, P < 0.0001, n = 9, Fig. 4.3b).
Calcareous-bound P was the largest sedimentary P pool in Shark Bay, comprising an
average of 54.8% of total P in Shark Bay sediments (Table 4.3). The concentration (R2 =
0.2, P < 0.0001, n = 144) and proportion (R2 = 0.66, P < 0.0001, n = 48) of calcareous
bound P in sediments significantly decreased with increasing salinity. Low salinity sites
had the highest concentrations and proportions of calcareous P, with a maximum of 60.5
μg P g -1 (69.6%) at site 2 (Table 4.2). High salinity sites had the lowest concentrations
and proportions of calcareous P in sediments. There was no relationship between
calcareous P concentrations and sediment depth (F5,138 = 0.25, P = 0.94), with mean
concentrations remaining relatively constant in the upper 40 cm of sediment (34.3 μg P g
DW-1 at 20-30 cm to 37.6 μg P g -1 at 5-10cm; Fig. 4.2).
81
Residual organic phosphorus (µg P g−1 )
Fraser (2016)
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750
Total sediment sulfur (mg/kg)
Figure 4.3: Scatterplot showing relationship between (a) sediment residual organic P concentrations and
sediment organic matter concentrations (R2 = 0.50, P < 0.0001, n = 144) and (b) total sedimentary sulfur
and proportion of residual organic P in sediments (R2 = 0.98, P < 0.0001, n = 9). Dashed line represents a
linear model added to each dataset; grey outline shows 95% confidence interval.
Solution 31P-NMR spectroscopy
Inorganic orthophosphate was the dominant P compound extracted from sediments by
NaOH-EDTA, and gave an 31P-NMR signal of δ=6.5ppm; however this signal shifted to
δ=4.5ppm at site 6, likely due to this extract containing higher concentrations of the
paramagnetic magnesium than other samples (Table 4.1). Inorganic phosphate extracted
by EDTA was higher at low and intermediate salinity sites, and decreased at higher
salinities. Phosphate monoesters were the only forms of organic P detected, which gave
82
Phosphorus cycling in seagrass sediments
signals between δ=4.5ppm and 5.0ppm (Fig. 4.4, inset). Phosphate monoester peaks were
greatest at intermediate salinities (sites 4 and 5), but could not be detected at sites 6 and
7 (the sites of highest salinity). Site 8 had relatively high monoester peaks, even though
the inorganic phosphate peak at this site was lower relative to lower salinity sites.
Figure 4.4: Solution 31P-NMR spectra from Shark Bay sediments by NaOH-EDTA extract. Figure inset
shows area of spectrum from site 5, where organic P in samples was detected, most likely as phosphate
monoesters. A line broadening of 5 Hz was used in the spectra.
Discussion
The concentrations and relative proportions of sediment P fractions changed substantially
across the salinity gradient in Shark Bay, although not exactly as we had expected. For
example, more labile fractions like magnesium-extractable P and reducible P did not
decrease in a predictable manner with increased salinity. However, we did find complex
shifts in the overall composition of sediment P that has significant implications for
primary productivity of seagrass meadows. At low salinity sites (e.g. sites 1-3, <40‰),
the bulk (>60%) of sedimentary P is calcareous bound P (Table 4.2, Fig. 4.2). This is
considered the least available P fraction in marine environments given that the pH
changes required for P release are unlikely to yield much P in buffered marine waters
83
Fraser (2016)
(Long et al. 2008). These low salinity sites also have the highest SRP (~0.2µM; Atkinson
1987) in the overlying water column, and is likely the dominant source of P for seagrasses
at these sites. At intermediate salinities (sites 4 and 5, 40-44‰), SRP concentrations in
the overlying water column are lower but labile P fractions like magnesium-extractable P
and reducible P fractions increase in the sediment. At these sites, root uptake is likely an
effective and more important strategy for P acquisition. The likely increasing importance
of sediment P for plant growth at sites where SRP is low is backed up by the strong decline
in magnesium extractable P with increasing sediment depth at these sites. At high salinity
sites (sites 6-8, >44‰), total sediment P and water SRP are at a minimum, yet the
proportion of organic P is highest at these sites. Therefore, recycling of sedimentary
organic P (primarily from seagrass detritus) and efficient uptake strategies would be
required at these high salinity sites. Given that salinity, sediment OM and SRP in the
water column all differ along the environmental gradient studied in this experiment, it is
not possible to identify if one variable in particular drives the fractionation of P in
sediments. We hypothesise that the combination of all factors combine to significantly
alter biogeochemistry along the gradient. Nevertheless, as salinity increases and water
SRP decreases in oligotrophic seagrass meadows, strategies to access sedimentary
organic P fractions become increasingly important to sustain primary productivity, with
the organic P fractions quantitatively increasing given higher OM in sediments.
The only organic P compounds detectable with
31
P-NMR analysis were phosphate
monoesters (Fig. 4.4). This is the first study that has sought to characterise the organic P
fraction in seagrass sediments and, the findings share some similarities but also have key
differences with studies of other terrestrial and aquatic ecosystems. Though phosphate
monoesters typically dominate the organic P pool in natural soils and both freshwater and
marine sediments, other organic P compounds are usually detected (Doolette and Smernik
2011; Baldwin 2013; Kraal et al. 2015; McLaren et al. 2015).For example, while
phosphate monoesters were the dominant organic P pool in Baltic Sea sediments, DNAP, techoic acid P and pyrophosphates were also detected (Ahlgren et al. 2006; Kraal et al.
2015). Similarly, phosphate monoesters dominate the organic P pool in mangrove
sediments, but phospholipids and phosphate diesters were also found (El-Rifai et al.
2008), with similar patterns seen in terrestrial (Doolette and Smernik 2011; Turner and
Engelbrecht 2011; Celi et al. 2013; McLaren et al. 2015) and wetland ecosystems (Turner
84
Phosphorus cycling in seagrass sediments
et al. 2007; Cheesman et al. 2012). Here the total organic P fraction comprised only 7-25
μg P g -1, and this low concentration accounts for the weak NMR signals. Regardless, the
lack of detection of readily degraded organic P compounds such as phosphate diesters
also indicate that if other organic compounds are present they are in very low
concentrations in Shark Bay sediments. Phosphate monoesters are typically the byproduct of plant tissues (Turner et al. 2002), and can be quite stable and thus persist longer
after deposition, requiring phosphatase enzymes to be released for hydrolysis to occur
before P was released (Ladd and Butler 1972; Burns 1982 Nannipieri 1998). Given that
OM in Shark Bay sediments is composed primarily of seagrass detritus (Cawley et al.
2012) and low SRP concentrations limit planktonic productivity (Kimmerer et al. 1985),
P compounds from seagrass litter likely form the bulk of the organic P pool in Shark Bay
sediments. We suspect that phosphate monoesters also dominate the sedimentary organic
P pool at other sites where the main sources of OM are seagrass detritus; further sampling
across a range of sites is needed to confirm this hypothesis.
Phosphate monoesters are essential signaling molecules present in all plants, commonly
released in response to stresses (Gillaspy 2011). For example, inositol phosphates have
been implicated in tolerance to increases in salinity in rice and tobacco (Majee et al. 2004;
Das-Chatterjee et al. 2006). A similar role in seagrasses may represent an adaptation that
allows the survival of these vascular plants in the saline marine environment, and may
partly explain the higher peaks in the NMR spectra from some of the higher salinity sites
(Fig. 4.4). However, concentrations of inositol compounds did not increase in Cymodocea
nodosa at sites with elevated salinities from the Mediterranean (Drew 1978), suggesting
that this stress response may be species-specific. The role of inositol phosphates for
salinity tolerance in seagrasses is largely unknown, yet may be a key physiological
response to salinity stress. Further
31
P-NMR analysis in different seagrass sediments
would show if the dominance of phosphate monoesters in the organic P pool of Shark
Bay sediments were a result of the increased salinities, the low concentrations of
phosphate available for plant uptake, or is a common feature across all seagrass
sediments. Such information would assist in determining the general mobility and
ecological importance of organic P in seagrass ecosystems.
The dominance of phosphate monoesters and lack of more readily degraded organic P
85
Fraser (2016)
compounds suggests that internal recycling of phosphorus is important in Shark Bay
sediments, which is typical of low P environments (Reed et al. 2011). To utilize the
organic P pool in Shark Bay, organisms must either produce enzymes that can hydrolyse
phosphate monoesters, or rapidly recycle and take up more labile organic P compounds
present at very low concentrations. Seagrasses and attached epiphytes release alkaline
phosphatase on leaf surfaces when P limited, alleviating P limitation by hydrolysing
phosphate monoesters (Pérez and Romero 1993; Martínez-Crego et al. 2006; Koch et al.
2009). However, the capacity for seagrass roots to release phosphatases into the
rhizosphere is largely unknown, despite the large proportion of P in organic forms in
sediments. Microbialogically derived phosphatases are also important in phosphate
monoester cycling in terrestrial soils (Richardson and Simpson 2011), and can be more
efficient than plant derived phosphatases (Tarafdar et al. 2001), but their ecological
importance in seagrass sediments is unknown. Activity of microbial and seagrass derived
phosphatases in sediments would likely be important in ecosystems like Shark Bay that
are characterised by low available P concentrations yet contain large organic P pools
dominated by phosphate monoesters that require hydrolysis before seagrass uptake.
Table 4.3: Summary table comparing P fractions in Shark Bay sediments to other seagrass sediments. Mean
sediment pools for magnesium extractable P (PMgCl), reducible P, calcareous P, and residual organic P are
provided, as well as dominant seagrass species and sediment OM concentrations. All P fractions are in μg
g-1. The proportion of each P fraction as total sedimentary P are also given below concentrations in
brackets. * indicates data was presented as calcareous + residual organic P combined.
86
Phosphorus cycling in seagrass sediments
The ephemeral Wooramel River has previously been hypothesised as a source of P for
seagrass uptake in Shark Bay after significant flow events, with inputs likely to be in the
reducible P fraction given the iron-dominated hinterland (Fourqurean et al. 2012; Fraser
et al. 2012). However, the Wooramel River had not flowed for over two years at the time
of sampling and we found no evidence of higher reducible P concentrations at the sites
adjacent to the Wooramel River (sites 6-8), even though iron and aluminium
concentrations were highest in these sediments (Table 4.1). While flood events in this
arid region are quite rare (once or twice a decade), the Wooramel River potentially can
introduce significant amounts of sediment to Shark Bay. Even though P concentrations
in Western Australian soils are low due to the highly-weathered nature of the continent,
rivers in semi-arid Australia still have the potential to bring large pulses of P after
rewetting following large rainfall events (Mitchell et al. 1997; McIntyre et al. 2009).
Concentrations of reducible P are significantly affected by changes in abiotic variables
such as pH and redox potential (Wang et al. 2013). Under anoxic conditions and oceanic
pH, P is likely to be desorbed from sediments quite quickly following large rainfall events
and released to the overlying water column, promoting release of P and removing it from
the sedimentary pool (Jordan et al. 2008). The high salinities at the sites would also
promote release of P from reducible sediments (Sundareshwar and Morris 1999). Total P
in sediments and seagrasses adjacent to the Wooramel River were elevated following the
2010/11 Wooramel River floods (Fraser et al. 2012), but this appears to be short-lived as
all labile fractions were low by the time of sampling for this study 24 months later.
However, given the low bioavailable P concentrations at the Wooramel Delta, such pulse
events may be important for seagrasses to periodically meet P demands, especially if the
P taken up by seagrasses can then be recycled into the organic P pool, preventing export
from the immediate area.
Overall, this study has revealed that Shark Bay may not be as oligotrophic and P-limited
as previously thought (Atkinson 1987; Burkholder et al. 2013). Shark Bay does indeed
have lower total sediment P concentrations in comparison to other similar embayments
(Table 4.3). However, we found that only 40-60% of total P is bound to calcareous P,
which his much less than other tropical seagrass ecosystems such as Bermuda, where up
to 77% is present in calcareous bound fractions. Proportions of labile P (magnesiumextractable P) are also relatively high in Shark Bay compared to similar embayments.
87
Fraser (2016)
However, Shark Bay has similar proportions of calcareous bound P (~50%) and organic
fractions to sediments in Florida Bay (16-40% in Shark Bay, ~40% in Florida Bay). The
relatively high organic P is likely almost solely derived from seagrass detritus in both
ecosystems (Koch et al. 2001; Cawley et al. 2012). Thus, the recycling of autochthonous
seagrass OM likely plays a key role in P cycling in sub-tropical oligotrophic embayments,
and becomes increasingly important as SRP decreases and salinity increases.
Acknowledgments
An NHT-II Caring for our Country project grant coordinated by Western Australian
Marine Science Institute awarded to G.A.K. funded this research. Additional funding for
M.W.F was received through PhD funding from the School of Plant Biology, UWA.
Logistical support was provided by the Shark Bay Ecosystem Research Project under the
direction of Michael Heithaus, funded by NSF (OCE-0745606). We thank Oscar SerranoGras, John Statton and Amy Griffiths for help collecting samples, and Kate Bowler for
assistance in sequential extractions. The authors acknowledge the facilities, and the
scientific and technical assistance of the Australian Microscopy & Microanalysis
Research Facility at the Centre for Microscopy, Characterisation & Analysis, The
University of Western Australia, a facility funded by the University, State and
Commonwealth Governments.
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language and environment for statistical computing.
Reed, S. C., A. R. Townsend, P. G. Taylor, and C. C. Cleveland. 2011. Phosphorus
Cycling in Tropical Forests Growing on Highly Weathered Soils, p. 339–369. In E.K.
Bünemann, A. Oberson, and E. Frossard [eds.], Phosphorus in Action. SpringerVerlag, Berlin.
Richardson, A. E., and R. J. Simpson. 2011. Soil microorganisms mediating phosphorus
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Ruttenberg, K. C. 1992. Development of a sequential extraction method for different
forms of phosphorus in marine sediments. Limnology and Oceanography 1460–1482.
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Solórzano, L., and J. Sharp. 1980. Determination of total dissolved phosphorus and
particulate phosphorus in natural waters. Limnology and Oceanography 25: 754–758.
Sundareshwar, P., and J. T. Morris. 1999. Phosphorus sorption characteristics of intertidal
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44: 1693–1701.
Tarafdar, J. C., R. S. Yadav, and S. C. Meena. 2001. Comparative efficiency of acid
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Turner, B. L., and B. M. J. Engelbrecht. 2011. Soil organic phosphorus in lowland tropical
rain forests. Biogeochemistry 103: 297–315. doi:10.1007/s10533-010-9466-x
Turner, B. L., E. Frossard, and D. S. Baldwin. 2005. Organic phosphorus in the
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Turner, B. L., M. J. Paphazy, P. M. Haygarth, and I. D. Mckelvie. 2002. Inositol
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Turner, B. L., N. Mahieu, and L. M. Condron. 2003. Phosphorus-31 Nuclear Magnetic
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Turner, B. L., S. Newman, A. W. Cheesman, and K. R. Reddy. 2007. Sample Pretreatment
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freeze/thaw episodes. Eur J Soil Science 45: 353–359.
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Chapter 5: Taxonomic and functional structure of microbial
communities in seagrass sediments across salinity and
phosphorus gradients
Matthew W. Fraser, Deirdre B. Gleeson, Pauline F. Grierson, Bonnie Laverock, Gary A.
Kendrick
Preamble: This manuscript is to be submitted to Nature Microbiology, and is formatted
to the requirements of the journal. The requirements of Nature Microbiology include
having an introductory paragraph with results and implications instead of having an
abstract, methods being presented after the main body of text, and no distinct results or
discussions sections. The work is primarily my own. DBG and BL provided advice
regarding microbial analysis. GAK and PFG provided advice regarding experimental
design. All co-authors provided feedback on the manuscript. This chapter examines the
shift in microbial community structure and function along the salinity/P availability
gradient in Shark Bay sediments described in previous chapters, and examines
correlations to changes in sediment biogeochemistry. I also compare the metagenomic
structure of microbial communities in seagrass sediments to other ecosystems to provide
context for the results.
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Sediment microorganisms influence primary producer composition and ecosystem
functions given their critical role in regulating biogeochemical processes. Sediment
microorganisms are particularly importantly in oligotrophic ecosystems where
small changes in the availability of macronutrients such as phosphorus (P) have the
greatest impacts. Seagrass meadows are marine ecosystems that present a unique
ecological niche to sediment microbial communities as seagrasses possess true root
systems that release oxygen and organic matter. Rhizosphere microorganisms
mediate nutrient cycling and organic matter turnover in seagrass sediments 1, but
the identity of the microbes and how they may respond to chemical change in their
aquatic environment is unknown. Here, we used metagenomic sequencing in
seagrass sediments in order to investigate the structure and functioning of the
microbial community across a salinity and P availability gradient in Shark Bay – an
oligotrophic, hypersaline, and arid ecosystem that boasts abundant seagrass
meadows that directly contribute to its World Heritage Status2. We show that
microbial communities change appreciably across salinity and P availability
gradients, with increases in phosphonate metabolism genes showing adaptations
toward increasingly P-limited conditions. Given that sediment organic P
concentrations are highest where P concentrations in the water column are low, the
microbial processing of organic P may support P requirements of seagrasses at
particularly oligotrophic sites.
Microorganisms have an integral role in maintaining productivity in oligotrophic
ecosystems by possessing genes that encode for novel enzymes that enable the uptake of
organic phosphorus pools3. Where nutrients become more limiting, genes that promote
efficient nutrient uptake and recycling increase in microbial communities, leading to
increases in the release of extracellular enzymes capable of degrading organic matter and
altering nutrient cycling rates4,5. For example, in the ultra-low P Sargasso Sea, novel
organic phosphorus compounds such as polyphosphate are used as the main P source,
driven by microorganisms that have an increased capacity for storing and degrading these
compounds6,7. Similarly, soil microorganisms often form close associations with plants
in terrestrial oligotrophic ecosystems, and are important in maintaining ecosystem
productivity given their ability to degrade refractory organic P forms, with biological
mineralization of organic P responsible for up to 90% of plant P uptake8. Seagrass
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Phosphorus cycling in seagrass sediments
meadows present unique/novel habitats for sediment microorganisms in a marine setting
as the only group of marine angiosperms. As angiosperms, seagrasses possess true root
systems and rhizospheres, oxygenating surrounding rhizosphere sediments and creating
unique conditions that support high levels of bacterial diversity9. Microbial communities
within the seagrass rhizosphere may also increase nutrient availability, especially in
oligotrophic ecosystems10. However, studies of microbial communities in seagrass
rhizosphere sediments have to date primarily focused on their taxonomic structure9,11,12,
with limited information on their functional capacity to influence biogeochemical
processes and ecosystem function in seagrass meadows13. This information is an
important knowledge gap given the numerous ecosystem services seagrass meadows
provide, including provision of habitat for other marine organisms including important
fisheries species14, improvement of water quality through nutrient uptake, and enrichment
of the organic matter content of sediments, contributing greatly to the global carbon cycle
as a key blue carbon habitat15.
Environmental gradients alter microbial community function and thus enzyme expression
and biogeochemical processing16. Salinity and nutrient availability are two key
environmental drivers in seagrass ecosystems, influencing productivity and trophic
structure17-19. Shark Bay has one of the largest seagrass meadows in the world, and
contains a natural salinity and phosphorus (P) availability gradient. Salinity rises from
normal marine salinity (~35‰) in the north of the Bay to over 65‰ in the southern,
hypersaline reaches of the Bay 2. Soluble reactive phosphorus (SRP) – the most available
source of P for biological uptake – decreases along this same gradient and is extremely
low (<0.02µM) in high salinity areas20,21. Seagrasses are extremely important in
maintaining the status of Shark Bay as a World Heritage Site, underpinning a diverse
ecosystem including turtle and dugong populations, and contributing to the salinity
gradient (by restricting hydrodynamic circulation) that supports Earth’s most extensive
stromatolite populations2. In addition, the seagrasses of Shark Bay sequester 243 Mg
organic carbon ha-1 in sediments, making it a global hotspot for coastal organic carbon
burial and comparable with organic C stored in terrestrial forests22. However, seagrassderived organic matter also influences biogeochemical processes and microbial
community structure in sediments23, and may be an important source of nutrients for
seagrasses in oligotrophic areas. It is unknown how the current environmental conditions
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within Shark Bay impact upon seagrass-associated microbial communities and the
processes they mediate, and how future environmental changes (e.g. increases in
temperature, increased salinities, changes in nutrient availability) may also influence
these microbial communities.
Here, we used metagenomic sequencing to determine taxonomic and functional changes
in the rhizosphere sediment microbial communities of 6 seagrass meadows in Shark Bay
(Fig. 5.1) encompassing a range of salinities (39.2‰ – 53.2‰, Appendix 1) and water
SRP concentrations from 0.1µM to below detectable concentrations (<0.02µM). We
hypothesised that: (i) rhizosphere microbial communities would significantly change in
taxonomy and function along the gradient; (ii) microbial communities at high salinity/low
SRP sites would possess functions that are advantageous for P uptake in such
environments (i.e. higher abundance of phosphorus metabolism genes); and (iii)
rhizosphere microbial function would be correlated to enzyme activities. We also
compared the microbial metagenomes of seagrass rhizospheres from this study with other
aquatic and terrestrial ecosystems to ascertain if there are consistent patterns in microbial
community structure and key biogeochemical parameters that would in turn indicate
potential “universality" in the adaptations of microbial communities to particular
conditions.
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Phosphorus cycling in seagrass sediments
Figure 5.1 Location of study sites in Shark Bay, Western Australia. Colored dots show sites where
metagenomic profiles were generated. White dots denote sites where additional samples where
biogeochemical measurements were also taken.
Across the salinity/P availability gradient, there were distinct shifts in the taxonomic and
functional structure of seagrass-associated microbial communities. Sites above 46‰
salinity showed separation in taxonomic profiles compared with profiles generated from
communities
sampled
at
lower
salinities
(Fig.
5.2a).
Proteobacteria
were
underrepresented in the high salinity sites (7 and 9, Fig. 5.2a). Gene abundances for
Betaproteobacteria (η2=0.64, P=0.015,), Deltaproteobacteria (η2= 0.57, P=0.029),
Gammaproteobacteria (η2=0.78, P=0.0023), and Zetaproteobacteria (η2= 0.88,
P=0.0002,) all significantly decreased along the salinity gradient. In contrast,
Alphaproteobacteria tended to increase with increasing salinity (η2= 0.53, P=0.038), as
did Planctomycetes (η2= 0.79, P=0.005) and Verrucomicrobia (η2= 0.85, P=0.001), with
highest abundances in sites 7 and 9 (Fig. 5.2b). The class Cytophagia (phylum =
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Fraser (2016)
Bacteriodes) also increased significantly with salinity (η2= 0.73, P=0.0059).
Functional gene abundances within the microbial community also shifted along the
salinity gradient and matched patterns in taxonomic composition, with sites above 46‰
differentiated from lower salinity sites (Fig. 5.3a, Appendix 2). Genes related to microbial
photosynthesis (P=0.008), DNA metabolism (P=0.01), and P metabolism (P=0.015) were
all significantly more abundant in high salinity sites, while nitrogen metabolism
(P=0.015), and cell signaling (P=0.044) were significantly lower. Clearly, changes in
environmental conditions (in this case, salinity and P availability) can alter both the
taxonomic and functional capacity of seagrass-associated microbial communities, and
these changes have the potential to fundamentally alter biogeochemical cycling of P, N,
and C (photosynthesis) in sediments.
The gradient in P availability also contributed to significant changes in the capacity of
microbial communities to utilize organic phosphorus compounds – in particular,
phosphonates. Phosphonates are a significant component of the dissolved organic P pool
in oceans24, and become an increasingly important P source to relieve P limitation in for
diazotrophs in oligotrophic ocean ecosystems25. Genes related to phosphonate
degradation, transport, and metabolism were all significantly more abundant at high
salinity sites where SRP concentrations decreased below detectable concentrations (Fig.
5.3b). Genes for phosphoacetaldehyde hydrolyse, important for phosphate release from
phosphonates, also showed greater abundance at higher salinities (Fig. 5.3b). The
increased abundance of phosphonate metabolism genes suggests an increased reliance of
microbes on organic P within the high salinity sites. Given that sediment organic P
increases in both concentration and as a proportion of total P as salinity increases in Shark
Bay, microorganisms with the functional traits that allow them to use this resource would
be at a competitive advantage. Phosphonates may therefore be an important source of
biological P uptake in oligotrophic coastal marine ecosystems as they are in open ocean
ecosystems.
98
Figure 5.2 Taxonomic profiles from metagenomic sequencing of microbial communities in Shark Bay seagrass sediments. (a) Principal component analysis
of taxonomic profiles between samples at the class level. Sites gradually increase in salinity from green (lowest salinity, site 1) to dark blue (highest salinity,
site 9). (b) Differences in relative abundance of classes between high salinity sites (>46‰, sites 7 and 9, teal) and lower salinity sites (<46‰, grey). Only
significantly different classes are shown, with corrected P-values calculated using Storey’s false discovery rate approach (P<0.05).
Phosphorus cycling in seagrass sediments
99
Figure 5.3 Functional profiles from metagenomic sequencing of microbial communities in Shark Bay seagrass sediments. (a) Principal component analysis of
functional profiles between samples at the class level. Sites gradually increase in salinity from green (lowest salinity, site 1) to dark blue (highest salinity, site 9).
(b) Differences in relative abundance of P metabolism functions annotated using the SEED Subsystems database between high salinity sites (>46‰, sites 7 and
9, teal) and lower salinity sites (<46‰, grey). Functions related to phosphonate metabolism are in bold. Only P-metabolism functions that were significantly
different between high and lower salinity sites are shown, with corrected P-values calculated using Storey’s false discovery rate approach (P<0.05)
Fraser (2016)
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Phosphorus cycling in seagrass sediments
Patterns in gene abundance do not always necessarily dictate actual expression of genes
and subsequent biogeochemical shifts26. In this study, gene abundance for secreted
alkaline phosphatase significantly increased with salinity (R2=0.98, P<0.0001) and were
strongly correlated with extracellular activity of the alkaline phosphatase enzyme
measured in sediments along the gradient (R2=0.88, P<0.0001). This suggests that the
alkaline phosphatase in the sediments is likely of microbial origin. In agreement to
previous studies20, there was, however, no corresponding increase in P content of seagrass
leaves (Appendix 1). Assuming enough substrate, the release of phosphatase enzymes
produced by bacteria would increase the availability of organic P to both microorganisms
and seagrasses. As such, the activities of microorganisms in sediments may assist
seagrasses in meeting P requirements in extremely oligotrophic areas such as our lowSRP sites; and may constitute an important ecological interaction between seagrasses and
rhizosphere microbial communities.
Genes related to P, nitrogen, and sulfur metabolism were more abundant in seagrassassociated microbial communities than in most other ecosystems (Fig. 5.5c). In particular,
genes related to organic S and P metabolism were highly abundant in seagrass
ecosystems. Given that seagrasses can take up nutrients from the water column through
their leaves27, the remineralization of organic matter in sediments has often been
overlooked as a potential nutrient source, especially given that organic matter in seagrass
sediments is often refractory. However, seagrasses can receive organic inputs from
terrestrial and oceanic sources, and high sedimentation rates ensure much of this material
is retained in seagrass sediments. In addition, seagrass sediments are often enriched in
seagrass detritus itself. The sediment organic pool is potentially a major source of
nutrients in seagrass ecosystems, and may play an important role in maintaining
productivity. If so, the functional diversity of seagrass-associated microbial communities
will play a central role in nutrient cycling, primary productivity, and the maintenance of
ecosystem services that seagrasses provide.
101
Figure 5.4 Taxonomic profiles from metagenomic sequencing of microbial communities from a range of different ecosystems. (a) Principal component analysis of
taxonomic profiles (level=class) between microbial communities from all ecosystems examined. Seagrass communities from this study shown in bright green. (b)
Principal component analysis of taxonomic profiles (level=class) from a subset of ecosystems, with freshwater and stromatolite communities removed to show finescale patterns. (c) Relative abundance of classes between seagrass microbial communities and microbial communities from other ecosystems. Only significantly
different classes are shown, with corrected P-values calculated using Storey’s false discovery rate approach (P<0.05).
Fraser (2016)
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Phosphorus cycling in seagrass sediments
Seagrass-associated microbial communities, somewhat unsurprisingly, showed the
strongest similarity to mangrove microbial communities when compared to profiles from
a range of different terrestrial and aquatic ecosystems (Fig. 5.5). Mangroves and seagrass
habitats share many environmental similarities; both are dominated by rooted
macrophytes growing in organic-rich, anoxic sediments submerged in saline water28.
Microbial communities from other marine benthic environments, such as salt marsh and
estuarine sediments, also showed significant similarity to those from seagrass sediments.
However, when comparing the functional profiles for the same communities, profiles for
seagrass rhizosphere communities showed greater overlap with functional profiles from
mangrove, estuarine, and even some terrestrial ecosystems than with salt marshes (Fig.
5.5). This suggests that the environmental factors driving microbial community structure
and functional capacity are to some extent independent of each other. Microbial
community structure in marine habitats can also be largely driven by communities in
adjacent habitats such as the surrounding water or sediments, with subsequent selection
based on functional capabilities29. Seagrasses have unique microbial communities with a
distinct taxonomic and functional structure, likely driven by a combination of
environmental and biological factors that are yet to be fully characterised.
In addition to genes regulating the P cycle, there were also significant shifts in taxonomic
composition of sediment microbial communities sampled. The bacterial phylum
Planctomycetes was more abundant within the high salinity sites of Shark Bay (Fig. 5.2b).
Planctomycetes also appeared more abundant in seagrass sediments compared with other
ecosystems (Fig. 5.4). The Planctomycetes can have significant impacts on nutrient
cycling. For example, some anaerobic Planctomycetes can perform anaerobic ammonia
oxidation, greatly altering N availability in anaerobic environments30, and are responsible
for up to 50% of fixed N removal from the ocean31. Planctomycetes also produce
abundant sulfatases, catalysing the hydrolysis of organosulfates and impacting the sulfur
cycle32. There has been recent focus on their interactions with macroalgae given the high
abundance of Planctomycetes on macroalgal biofilms in comparison to their negligible
abundance in seawater33-35. The interaction between some macroalgal species and
Planctomycetes is thought to be symbiotic; with the Planctomycetes contributing to the
remineralization of organic matter and mobilization of nutrients, in return for habitat and
substrate provision by the macroalgae33. Planctomycetes have previously been found in
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other seagrass sediments9-11,36, but their specific interactions with seagrasses are
unknown. Planctomycetes abundances may be high in seagrass sediments because of
seagrass-derived organic exudates or litter23,37, providing an ideal niche where they can
utilize a range of chemical compounds for energy33. Seagrasses may also benefit from
this association by making use of the nutrients released by Planctomycetes metabolic
processes. If so, Planctomycetes may be key organisms mediating the turnover of organic
matter in seagrass ecosystems, becoming increasingly important as nutrient availability
decreases.
104
Figure 5.5 Functional profiles from metagenomic sequencing of microbial communities from a range of different ecosystems. (a) Principal component analysis of
functional profiles (level 1 functions) between microbial communities from all ecosystems examined. Seagrass communities from this study shown in bright green.
(b) Principal component analysis of functional profiles (level 1 functions) from a subset of ecosystems, with coral, freshwater and stromatolite communities removed
to show fine-scale patterns. (c) Differences in relative abundance of functions between seagrass microbial communities (green) and microbial communities from other
ecosystems (dark grey). Only significantly different classes are shown, with corrected P-values calculated using Storey’s false discovery rate approach (P<0.05).
Phosphorus cycling in seagrass sediments
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Methods
Sampling sites and experimental design
Rhizosphere sediments and plants were collected from nine shallow (1.5 – 2.5 m) subtidal
sites across Shark Bay, Western Australia in March 2014. Shark Bay is a World Heritage
listed embayment, with the seagrasses in the Bay crucial to this status2. The seagrasses of
Shark Bay support a diverse food web, are highly productive, and species rich. The
seagrass communities at all sites were dominated by the temperate seagrass Amphibolis
antarctica, with the tropical species Halodule uninervis as a sparse understory.
The nine sites were chosen to encompass a range of salinities and SRP concentrations,
from site 1 (lowest salinity, Guichenault Point) to site 9 (highest salinity, L’haridon Bight)
(Fig. 5.1). All sites were separated by at least 5km. At each site, four sediment samples
(0-10 cm deep) were taken using 50 mL plastic syringe core, and transferred to 50 mL
centrifuge tubes, for later biogeochemical and enzyme analysis. Amphibolis antarctica
leaves (n=100) were also hand-collected at each site for P content analysis. Samples were
stored on ice in the field for no longer than 8 hours, and subsequently stored at 4°C
(sediment samples) or -20°C (seagrass leaves) until processing. In addition, five extra
sediment samples (0-5 cm) were collected using an ethanol-sterilized 3ml syringe core,
transferred to 3 mL cryovials and snap frozen immediately in liquid nitrogen. Cryovials
were then stored at -80°C until DNA extraction. On return to the laboratory, seagrass
leaves were thawed and epiphytes were removed from leaves by gentle scraping with a
razor blade under deionised water. Leaves were dried at 60°C for 3 days, then ground
using a mortar and pestle. P content on a subsample was determined using standard dry
oxidation acid hydrolysis techniques38. SRP for water samples was measured
colorimetrically on filtered (0.45 µm) water samples collected from each site using
standard methods39.
Enzymes
The activities of three hydrolytic enzymes were measured using standard colorimetric
methods40. The enzymes and substrates used were (i) β-glucosidase assayed with 4nitrophenyl β-glucopyranoside, (ii) acid phosphatase and (iii) alkaline phosphatase both
assayed with 4-nitrophenyl phosphate disodium hexahydrate. For each assay, 1.00 g of
sediment was added to a 50 mL Erlenmeyer flasks with 4 mL of modified universal buffer
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Phosphorus cycling in seagrass sediments
(pH=6.00 for β-glucosidase, pH=6.50 for acid phosphatase and pH=11.00 for alkaline
phosphatase) and 1 mL substrate, and were incubated at 37 ºC for one hour. Following
incubation, 1 mL of 0.5M CaCl2 and 4 mL of 0.5M NaOH (for phosphatases) or 4 mL
10mM tris(hydroxymethyl) aminomethane (pH=12, for glucosidases) were added to stop
the reaction. The final nitrophenol concentrations were determined photometrically at 400
nm against a standard curve. Microbial P was operationally calculated as the difference
in resin-extractable P between non-fumigated and hexanol-fumigated sediment samples
using standard methodology41.
Microbial community identification
Differences in sediment microbial communities were firstly identified across all nine sites
using automated ribosomal intergenic spacer analysis (ARISA), before a subset of
samples was chosen for further metagenomic analysis. A subset of six sites was selected
for metagenomic analysis (sites 1, 3, 5, 6, 7 and 9). DNA extraction from 500 mg of three
replicate sediments per site was performed using PowerSoil DNA Isolation Kit (MoBio,
California, CA, USA) according to the manufacturer’s instructions. Where extractions
yielded less than 200 ng of DNA, additional extractions were performed to ensure final
concentrations of neat DNA were above 200 ng. In total, 18 DNA samples were
sequenced (Australian Genome Research Facility, Australia) on an Ilumina MiSeq
sequencer. Two runs were performed to ensure that a target of 2 million sequence reads
per sample.
Metagenomic annotation and statistical analysis
Raw, unassembled reads were annotated using the Metagenomics Rapid Annotation
(MG-RAST) pipeline version 3.642. Taxonomic profiles were generated using
Representative Hit Classification of the M5NR database to allow comparisons of multiple
metagenomes42, and functional profiles were generated using Hierarchical Classification
against the Subsystems database, with a minimum alignment length of 50 bp and E value
cutoff of E < 1x10-5 43,44(Appendix 3).
To examine statistical differences in taxonomic and functional profiles across Shark Bay
sediments, data were analysed using the Statistical Analysis of Metagenomic Profiles
(STAMP) package45. P values for all tests were calculated using ANOVA for multiple
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test comparisons and Tukey-Kramer test for posthoc analysis. Storey’s false discovery
rate (FDR) was used for multiple test corrections, given its higher power than other
correction methods such as Benjamini-Hochberg FDR46,47. Effect sizes (η2) are provided
with P values to show magnitude of shifts and give context to the biological significance
of significant results48.
To compare microbial communities from seagrass rhizosphere sediments in this study to
those from other ecosystems we downloaded metagenomic profiles from a range of
terrestrial, aquatic, and marine habitats publicly available on MG-RAST (full details
Appendix 4). Where possible, we chose profiles obtained with similar average sequence
lengths (~300 bp) and using Ilumina as a sequencing platform, though this was not always
possible. Again, taxonomic profiles were generated using Representative Hit
Classification of the M5NR database, and functional profiles were generated using
Hierarchical Classification against the Subsystems database, with a minimum alignment
length of 50 bp and E value cutoff of E<1x10-5 43,44. Profiles were combined and analysed
using STAMP, with post-hoc tests using the Games-Howell test to account for the
unequal sample sizes45. Storey’s FDR was used for multiple test corrections.
Relationships between sediment enzyme activity, microbial P, sediment organic matter
content, and selected metagenomic groups were analysed using linear regression analysis
against salinity. Prior to analysis, data were checked for normality and homogeneity of
variances. Due to assumptions being violated, non-parametric tests (Kruskal Wallis) were
used to examine the relationships for seagrass P content. Statistical tests were analysed
using R version 3.1.148 using the base program and the ‘ggplot2’ package for graphics49.
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Chapter 6: Tracking phosphorus cycling in seagrass meadows
using oxygen isotopes: Methodological challenges and future
approaches
Matthew W. Fraser, Pauline F. Grierson, Gary A. Kendrick, Grzegorz Skrzypek
Preamble: This chapter describes an ultimately unsuccessful attempt to investigate the
extent to which microbial turnover of organic P and subsequent uptake of mineralised P
by seagrass using stable oxygen isotope composition of phosphates (δ18O-PO4) as a tracer
for biological cycling of P. An analysis of the approach taken and some of the challenges
faced in this research is included in the thesis in order to inform future efforts of δ18OPO4 measurements in seagrass ecosystems.
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Introduction
In oligotrophic seagrass meadows, concentrations of soluble reactive phosphate (SRP) in
the water column may be too low to sustain productivity, and uptake of P from sediment
sources become increasingly important in meeting seagrass P demands (Chapters 4 and
5). While the effect of geochemical processes such as sorption or desorption on P
availability in marine sediments are relatively well understood (Paytan & McLaughlin
2007), less is known about the importance of the ecological and biological mechanisms
of P regeneration from microbial biomass. In oligotrophic meadows with low inorganic
P concentrations, mineralization of sediment organic P compounds including the
mineralization of the microbial biomass may be important in sustaining P requirements
of seagrasses.
The isotopic composition of oxygen (δ18O) in soluble reactive phosphorus (SRP) is
thought to reveal the balance between phosphate transport and biological turnover rates
in marine ecosystems (Colman et al. 2005). Tracking the importance of biological P
cycling is more difficult than carbon or nitrogen cycling given that P only has one
naturally occurring stable isotope (31P). Radioisotopes (32P and
33
P) can be used to
measure biological P cycling (Van Veen et al. 1987; Benitez-Nelson & Karl 2002;
Nielsen et al. 2006), but only in short term studies that are often cost-prohibitive.
Recently, isotopes of oxygen on phosphate molecules have been recognized as a potential
method for tracking sources and cycling of P in aquatic environments (Paytan &
McLaughlin 2011). The P-O bond in phosphate is resistant to inorganic hydrolysis at most
temperatures on Earth, and isotope exchange of O can only occur during biological
processes, when δ18O-PO4 is equilibrated with water available for the organism, making
it is an excellent signature for biological activity (Tudge 1960; Blake, Alt & Martini
2001). Two processes control the δ18O-PO4 signature of plants or microorganisms. Direct
uptake of surrounding inorganic phosphate results in a δ18O-PO4 signature equal to
surrounding water/porewater, and shows a negligible contribution of P from organic
sources. Conversely, the extracellular degradation of organic P (and subsequent uptake)
results in δ18O-PO4 values lower than the the δ18O-PO4 of surrounding water (Blake,
O'Neil & Surkov 2005). The most crucial processes in this exchange is typically the
hydrolysis of phosphoesters (Blake et al. 2005; Liang & Blake 2009), which often
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Fraser (2016)
constitute the majority of organic P compounds. Given the capacity for the δ18O-PO4 to
indicate use of organic P compounds, Tamburini et al. (2012) used δ18O-PO4
measurements from sediments, vegetation, and microbial pools to determine that P is
cycled through soil microorganisms before being released to the available pool. By
comparing δ18O-PO4 signatures from all potential sources to the available P signature, it
may be possible to determine key processes leading to P becoming available for plant
uptake.
Previously, I have shown that organic P in the sediments of Shark Bay is dominated by
phosphate monoesters (Chapter 4). Soluble reactive P in the water column tends to be
negatively correlated to organic P in sediments. The expression of extracellular
phosphatase enzymes also increases along the SRP gradient, and microorganisms have
increased abundance of genes for the metabolism of organic P compounds at low SRP
sites (Chapter 5). Viewed collectively, these findings suggest that the microbial-driven
mineralization of organic P is an increasingly important P source for seagrasses along the
SRP gradient in Shark Bay. Here, we aimed to track biological turnover of phosphates by
measuring sediment, microbial, and seagrass δ18O-PO4 pools across the SRP gradient in
Shark Bay. We expected that seagrasses at the most oligotrophic (i.e. low SRP) sites
would show a δ18O-PO4 isotopic signature closest to the microbial P pool, reflecting an
increased uptake of P that had previously been recycled through sediment
microorganisms.
Methods
Sampling sites and experimental design
Sediment and seagrass samples were selected at nine shallow sites (1.5 – 2.5m) across
Shark Bay, Western Australia in March 2014 as described in Chapter 5 (see Fig. 5.1 for
full study map). The nine sites were chosen to encompass a range of salinities and SRP
concentrations, from site 1 (lowest salinity, Guichenault Point) to site 9 (highest salinity,
L’haridon Bight), with all sites separated by at least 5 km. The seagrass communities at
all sites were dominated by the temperate seagrass Amphibolis antarctica, with the
tropical species Halodule uninervis as a sparse understory. At each site, five sediment
samples (0-10 cm deep) were taken using a 50 mL plastic syringe core, and transferred to
50 mL centrifuge tubes, for later isotopic analysis of the sediment, microbial, and resin
114
Phosphorus cycling in seagrass sediments
extractable δ18O-PO4 pools. Amphibolis antarctica leaves (n=100) were also handcollected at each site for P content analysis. Samples were stored on ice in the field for
no longer than 8 hours, and subsequently stored at 4°C (sediment samples) or -20°C
(seagrass leaves) until processing.
Phosphorus extraction
Phosphorus was extracted from plants and sediments using the methodology of
Tamburini et al. (2012). Phosphate from three different sediment pools (resin extractable
P, microbial P, and calcareous sediment P) were extracted. Resin and microbial P
fractions were extracted as described by (Grierson & Adams 2000) in a modification of
the method of (Kouno, Tuchiya & Ando 1995). Resin extractable P represents P that is
readily desorbed from sediment particles and is considered available for plant uptake. For
resin extractable P, 20 g of fresh sediment was shaken with 4 cm2 strips of anion exchange
membranes (Part #BDH-551642S from VWR International ®, USA) in a 1:4
sediment:water suspension for 16h. Membranes were saturated with HCO32- prior to
extraction. Membranes were then removed from the sediment:water suspension, and
extracted phosphate was eluted from membranes by shaking the membranes in 0.2M
HNO3 for 30 minutes. Microbial P was also extracted using 20g of fresh sediment with
anion exchange membranes, but with 1 mL of hexanol added. Microbial P is thereafter
defined as the amount of extracted by the resin in the presence of hexanol minus resinextractable P without hexanol. Calcareous sediment P was extracted by adding 40 mL 1M
HCl to 10 g of sediment. Extracts were shaken overnight, filtered, and P was extracted
from the filtrate. This extraction was chosen as it is considered a good representation of
P present in Ca-P minerals (Tamburini, Bernasconi & Angert 2010). Foliar P of
seagrasses was extracted from 2 g of seagrass leaf material using a concentrated (10 M)
HNO3 extraction (Pfahler et al. 2012). Prior to extraction, seagrass leaves were scraped
clean of epiphytes, dried at 30°C for 48 hours, and ground using a mortar and pestle.
Phosphorus from phosphate rock standards (Standard NBS120c, National Institute of
Standards and Technology, USA) were also extracted using the same procedure as
calcareous P, and analysed to check P recovery.
Sample preparation and isotopic analysis
Phosphorus extracts from sediment and seagrass as well as reference material of known
115
Fraser (2016)
phosphate concentrations were purified and converted to silver phosphate in order to
remove any other compounds from samples that could contain oxygen that would
interfere with the analysis (Tamburini et al. 2010; 2012). Dissolved organic matter was
removed from P extracts using DAX-8 Amberlite Resin that had been preconditioned
using 100% methanol. Extracts were then precipitated as ammonium phospho-molybdate
(APM) by adding 4.2M NH4NO3 and 0.1M (NH4)Mo7O244H2O (ammonium
molybdate). APM crystals were left to form overnight, before being dissolved in citric
acid-NH4OH solution (10 g citric acid, 140 ml conc. NH4OH, 300 ml deionised water).
Extracts were then precipitated as magnesium ammonium phosphate (MAP) by adding
25 ml of a magnesia solution (50 g MgCl26H2O and 100 g NH4Cl dissolved in 1 L
deinoinsed water acidified to pH 1 with 12 M HCl), ammonia hydroxide solution (1:1 v:v
NH4OH/H2O) and leaving overnight. The MAP crystals were then thoroughly rinsed with
ammonia hydroxide solution, to remove traces of Cl that would interfere with later silver
phosphate precipitation. The MAP crystals were then dissolved in 0.5 M HNO3. The
resulting solution was then shaken overnight with BioRad AG50-X8 cation resin (H+
form, 100-200 mesh, preconditioned with HNO3) to remove cations from solution.
Finally, the samples were precipitated as Ag3PO4 by adding a Ag-amine solution (10.2 g
AgNO3, 9.6 g NH4NO3, 18.5 ml of concentrated NH4OH, and 81.5 ml deionised water).
Crystals were left to form overnight, before being vacuum filtered, rinsed with deionised
water, and left to dry at 50°C overnight. Typically, Ag3PO4 crystals should be yellow and
euhedral (Tamburini et al. 2010).
Extracted Ag3PO4 was weighed into silver capsules, and then pyrolysed in the presence
of carbon at 1450ºC in a helium stream in a TC/EA (Thermal Conversion/ Elemental
Analyser). After separation from other gases on a GC column, the yield CO was analysed
on a Delta XL Mass Spectrometer in continuous flow mode (Thermo-Fisher Scientific),
at the West Australian Biogeochemistry Centre. All δ18O-PO4 ratios are reported relative
to the VSMOW standard in per mil (‰) (Halas et al. 2011). Normalization to the
VSMOW scale was based on two replicated international reference materials provided by
the International Atomic Energy Agency (IAEA): IAEA-601, IAEA-602 following multipoint normalization (Skrzypek & Sadler 2011). The analytical combined uncertainty was
<0.4‰.
116
Phosphorus cycling in seagrass sediments
Results
Using standard extraction procedures, we successfully precipitated Ag3PO4 crystals from
phosphate rock standards. Ag3PO4 crystals from phosphate rock standards were yellow
and euhedral, consistent with previously successful experiments extracting Ag3PO4 from
sediments (Tamburini et al. 2010). Precipitate was also formed from calcareous sediment
and seagrass extracts, but had a very different appearance to the precipitates prepared
from the rock phosphate standards. Crystals from seagrass and calcareous sediment
extracts were brown-black, very fine, and flocculent. The appearance of the calcareous
sediment and seagrass precipitates was consistent across several extractions. Despite
repeated attempts, no precipitate was yielded from microbial or resin-extractable P
extracts.
Precipitates from standards, seagrass, and mineral P extractions were then analysed on
the isotope ratio mass spectrometer (IRMS). The δ18O-PO4 values from control standard
NBS120c were matching the commonly accepted values within typical for TC/EA
method precision. (Table 6.1). The CO yields for NBS120c where in the expected range
for pure AgPO4.
However, the signals for precipitates from seagrass and calcareous sediment P pools were
ten times lower than NBS120c standards and expected values had pure Ag3PO4 had been
extracted, and in several cases just above the background level typical for TC/EA. These
extremely low yields suggest that samples contained very little Ag3PO4, and likely
contained other unexpected low-oxygen chemical compounds that had been coprecipitated. Therefore, these results were discarded as not reliable.
Discussion
Given logistical and time constraints, I was unable to satisfactorily refine the δ18O-PO4
methodology to address the aims of this chapter. Though precipitates were extracted from
sediment and seagrass P pools, the low O2 yields prevent any confidence that the
extraction process was completed and that the extract contained pure Ag3PO4. The
successful precipitation and analysis of the phosphate rock NBS120c standard suggests
that the laboratory procedure was not the main cause for the unreliable results, and instead
117
Fraser (2016)
points towards problems with processing samples through current methods. Below, I
outline several potential reasons why the δ18O-PO4 analysis of Shark Bay sediments and
seagrasses was not successful and suggest future steps that may help overcome these
analytical issues.
Table 6.1: Maximum amplitude of mass 28 signal sampled (i.e. CO2 yield, mV) and δ18O-PO4 signatures
of Ag3PO4 precipitates from rock phosphate standards, seagrasses, and sediments. * - Oxygen yield too low
on samples, indicating presence of compounds other than Ag3PO4.
Sample
Sample
mass (g)
Rock
phosphate
standard
0.0005
-0.01
Phosphorus
concentration P content of Mass 28 CO2
(mg g-1)
sample (mg) yield (mV) δ18O-PO4 (‰)
21.4-22.8
(0.3)
(std. accepted
330
0.17 - 3.3
2321 - 2370 value=21.7)
Seagrass
Site 1
Site 2
Site 3
Site 4
Site 5
Site 6
Site 7
Site 8
Site 9
2
2
2
2
2
2
2
2
2
0.63
0.51
0.5
0.48
0.52
0.74
0.53
0.57
0.65
1.26
1.02
1
0.96
1.04
1.48
1.06
1.14
1.3
282
257
346
328
228
530
471
408
1102
19.2*
21.7*
20.2*
20.8*
21.3*
22.1*
23.1*
24.0*
22.4*
Sediment
Site 1
Site 4
Site 5
Site 6
Site 7
20
20
20
20
20
0.056
0.057
0.046
0.051
0.074
1.12
1.14
0.91
1.02
1.48
358
609
376
271
190
24.6*
14.2*
17.5*
4.4*
11.5*
Compounds within seagrasses that inhibit precipitation processes
The procedures used here have been successfully used to measure δ18O-PO4 in terrestrial
plants and surrounding soils (Tamburini et al. 2012; Pfahler et al. 2012) as well as marine
sediments (McLaughlin et al. 2006; Goldhammer et al. 2011b). However, to my
knowledge this is the first attempt to assess the δ18O-PO4 of seagrasses or their surface
sediments. Given the challenges faced here, and the appearance of the precipitates, it may
118
Phosphorus cycling in seagrass sediments
be that seagrasses contain compounds that cause contamination or interfere with the
formation of a pure Ag3PO4 precipitate. For example, seagrasses contain abundant
sulfated phenolic compounds to assist survival in a saline habitat and to deter herbivores
(McMillan, Zapata & Escobar 1980; Vergés et al. 2007). The concentration of these types
of compounds is rarely associated with other marine or terrestrial communities where
δ18O-PO4 has previously been measured. However, several precipitation and rinsing steps
are included in the published methodology which are designed to remove ions that could
potentially interfere with Ag3PO4 precipitation or δ18O-PO4 signals (Tamburini et al.
2010). Similarly, samples are initially treated with DAX-8 resin to remove dissolved
organic matter. The DAX-8 resin is highly effective at removing aquatic humic OM
compounds, but has primarily been used in freshwater environments, with limited use in
seagrass ecosystems (Lavery et al. 2013). We might speculate that organic compounds
that are more abundant in seagrasses may not be completely removed by this step, and
could have interfered with the precipitation of Ag3PO4. Further investigation utilizing
multiple methods for dissolved organic matter removal, coupled with spectroscopic
analysis of extracts, would help to determine how effective the DAX-8 resin is at cleaning
the samples of dissolved organic matter and whether an alternative procedure may be
required.
Hypersalinity of Shark Bay
Presence of ions such as Na+ and Cl- can limit the precipitation of Ag3PO4 (Tamburini et
al. 2014). As such, the high salinities in Shark Bay could be contributing to the limited
precipitation of Ag3PO4 in Shark Bay samples by concentrating specific ions that interfere
with the methodology. However, the methods used in this experiment have also been used
successfully in marine ecosystems (Colman et al. 2005; Goldhammer et al. 2011b; a;
Joshi, Kukkadapu & Burdige 2015) where salinities would also have been over 20‰. The
various precipitations used prior to the final Ag3PO4 precipitation include several rinsing
steps, and are designed to eliminate interfering ions. As such, salinity alone is unlikely to
have prevented proper Ag3PO4 precipitation. However, in hypersaline environments like
Shark Bay, it may be required to increase the number of rinses of precipitated crystals at
all stages of the method to account for the higher concentrations of anions and cations.
Cerium phosphate precipitations have been used in some marine δ18O-PO4 experiments
as an alternative to the APM precipitation (McLaughlin et al. 2004; 2006; Paytan &
119
Fraser (2016)
McLaughlin 2007). I also tried cerium phosphate extractions in place of APM
precipitations in this experiment, but again failed to purify Ag3PO4 from extracts.
Low phosphate concentrations
Both the waters of Shark Bay and the underlying sediments are highly oligotrophic. Thus,
there may have been insufficient phosphate in the sediment extracts for successful
precipitation of Ag3PO4. The purification of extracts into Ag3PO4 (with several different
precipitation and rinsing steps) likely leads to loss of phosphates, exacerbating the already
low concentrations. However, low concentrations (0.5 µmol) of porewater P have
successfully been extracted using similar methodologies, allowing ~200 µg Ag3PO4 that
was enough to generate adequate δ18O-PO4 signals (Goldhammer et al. 2011b). The
microbial and resin P extracts here were close to this lower limit of analysis; we suspect
loss of P through experimental handling likely left P concentrations too low. However,
calcareous sediment P and phosphates extracted from seagrass leaves were all above 1000
µg (Table 6.1), well in excess of the 200 µg limit successfully measured by Goldhammer
et al. (2011b). Overall low phosphate concentrations alone are unlikely to fully explain
the lack of Ag3PO4 precipitation.
Resin preparation
Goldhammer (2011) noted that using cation resin membranes prepared the previous day
resulted in reddish discolouration of samples that made proper Ag3PO4 impossible. The
reason for this discolouration was never fully determined, but was prevented by preparing
resin membranes immediately before analysis (Goldhammer, pers. comm.). I prepared
our resin immediately before use but despite this caution noticed a reddish discolouration
in some samples similar to that described by Goldhammer (2011). This may have have
contributed to improper Ag3PO4 precipitation but at this point I have no clear chemistrybased explanation as to why.
Final comment
My original objective for this study was to analyse the δ18O-PO4 signatures of different P
pools in seagrass ecosystems to elucidate biological P cycling in these environments.
However, I was unable to adequately extract purified Ag3PO4 from seagrasses or their
surrounding sediments. I propose that the poor precipitation of Ag3PO4 in this
120
Phosphorus cycling in seagrass sediments
experiments may be attributable to a combination of unique organic and sulfur rich
compounds in seagrass tissues, high salinities, and low P concentrations. Further
experiments should focus on determining the compounds in seagrass tissues that may
interfere with current analytical procedure, and find a method to remedy this (e.g. removal
of compounds in separate step, or new precipitation methodology). A first step would be
to spike samples from seagrass ecosystems with rock phosphate and test relative recovery.
It would also be useful to perform this experiment in an ecosystem with higher ambient
P concentrations and lower salinities, to avoid potential problems with high salinities and
low P concentrations. Systematic concentration-precipitation experiments need to be
undertaken to identify detection limits across a range of systems. It might also useful to
test procedures for concentrating P extracts prior to the precipitation step. New and
modified methods that are rigorously tested across a diversity of ecosystems and sample
types are thus needed in order to increase confidence in the δ18O-PO4 approach to
understand biological turnover of phosphates.
121
Fraser (2016)
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Chapter 7: General discussion: organic matter cycling and the
phosphorus biogeochemistry of Shark Bay
The findings of this thesis expand the knowledge of the different sources and significance
of organic matter in seagrass sediments, and the role OM plays in influencing P
biogeochemistry and sustaining the P demands of seagrasses in Shark Bay. More
generally, this research also provides new understanding of the ways in which seagrass
productivity is maintained in oligotrophic environments. This research also contributes to
the growing recognition of the significance of microbial communities and their
functioning to understanding the biogeochemical processes that sustain life in near coastal
ecosystems around the world. In particular, this research emphasizes the importance of
understanding the functional capabilities of rhizosphere microbial communities, given the
influence this has on key biogeochemical processes and ultimately primary productivity.
The primary aim of this PhD study was to characterise the sources and ecological impacts
of organic matter (OM) inputs into seagrass sediments, and then examine the role that
microbially-driven breakdown and turnover of this OM plays in supplementing
phosphorus (P) availability in an oligotrophic seagrass ecosystem (Figure 7.1). My
research demonstrates that OM plays a significant role in the P cycle of seagrass
meadows, and may supplement seagrass productivity in oligotrophic areas after microbial
mineralization. In this concluding discussion, I first briefly summarise the main findings
from my research. I then discuss how my research fits into wider paradigms emerging
within marine ecology more broadly, focusing on extreme climatic events, organic matter
cycling, and microbial ecology. I then discuss how my research changes our knowledge
of the biogeochemistry of Shark Bay, before concluding.
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Figure 7.1 Conceptual model showing main results from this thesis. 1) Seagrasses showed short term defoliation and longer-term loss of belowground
biomass in response to the 2010/11 marine heatwave, with biomass losses greatest in areas where light availability was reduced as a result of the 2010/11
Wooramel River flood, leading to an input of seagrass-derived OM into sediments. 2) Seagrass-derived OM into sediments led to shifts in nutrient availability,
enzyme expression, and microbial community structure in sediments, while seagrass seedlings has shifts in biomass allocation. 3) Sediment P fractionation
across the salinity and phosphate gradient in Shark Bay. High salinity sites had higher concentrations and proportions of organic P compounds, with phosphate
monoesters dominating this fraction. 4) Microbial communities showed an increase in the abundance of organic P at high salinity/low SRP sites, correlating
with an increase in the expression of phosphatases that could hydrolyse phosphate monoesters. 5) The microbial degradation could support the P requirements
of seagrasses in the sites with lowest SRP concentrations, but this could not be experimentally confirmed (using δ18O-PO4).
Fraser (2016)
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Summary of main findings of this research
I show that extreme climatic events that cause seagrass dieback lead to declines in
seagrass belowground biomass up to 2 years after disturbance, with high quantities of
seagrass-derived OM entering sediments (Chapter 2). Seagrass-derived OM significantly
changes belowground conditions in seagrass sediments; altering nutrient availability,
microbial community composition, and hydrolytic enzyme expression (Chapter 3), and
could be a major factor dictating sediment biogeochemistry in seagrass sediments during
seagrass colonization, meadow growth, and following dieback events. Further, OM in
sediments was found to be an important component of the P cycle in Shark Bay. As
salinity increased along the salinity gradient and soluble reactive phosphate (SRP) in the
water column decreased, the concentration and proportion of P in the organic P fraction
increased in sediments, even though total P decreased (Chapter 4). The organic P fraction
was dominated by phosphate monoesters, likely resulting from seagrass litter dominating
OM inputs into sediments. Microbial community structure and function in sediments also
significantly changed along the salinity/SRP gradient possibly in response to increased
sedimentary organic P (Chapter 5). Microbial communities had higher abundances of
genes related to the metabolism of organic P compounds at sites with high salinity/low
SRP sites, and were strongly correlated with an increased expression of phosphatase
enzymes that could hydrolyse phosphate monoesters. Though it is likely that the
microbial degradation of organic P at least partially supplements seagrass P requirements
in sites with low SRP, this could not be experimentally validated using δ18OPO4 as a tracer
of P sources due to methodological issues (Chapter 6). Nevertheless, this thesis clearly
show that seagrass-derived OM directly influences P biogeochemistry and microbial
communities within seagrass ecosystems. This thesis reveals new insights into the
significance of seagrass-derived OM, the characteristics of the organic P pool, and the
functional characteristics in seagrass sediments.
Sediment organic matter in seagrass meadows
Organic matter in seagrass sediments has recently received attention as a globally
important carbon sink, often noted for its refractory nature (Mcleod et al. 2011;
Fourqurean et al. 2012a). However, this thesis shows that OM cannot be considered as
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just a passive C sink in seagrass ecosystems, instead influencing biogeochemical
processes, microbial community structure, and seagrass physiology (Chapter 3-5). As
such, the gradual enrichment of sediments with OM after seagrass colonization may play
an important role in biogeochemical cycling and ecosystem processes in coastal
ecosystems. Furthermore, seagrass-derived OM contributes to detrital food webs that are
often more important to energy transfer up the food chain than direct grazing on
seagrasses, showing the bottom-up impacts this resource can have (Mcleod et al. 2011).
Shark Bay itself has been stated as a global C sequestration hotspot (Fourqurean et al.
2012b), but this sediment OM pool also has a contemporary ecological importance in
directly influencing biogeochemistry and productivity within Shark Bay.
The ecological and geochemical impacts of OM in seagrass sediments will largely depend
on the types of inputs the meadows receives. Oligotrophic ecosystems with abundant
seagrass meadows like Shark Bay have sediment OM pools dominated by seagrass
detritus characterised with high C:N and C:P ratios (Gacia, Duarte & Middelburg 2002;
Fourqurean et al. 2012b; Fraser et al. 2012). Conversely, settling organic material from
planktonic sources would make up a substantial component of total sedimentary OM in
mesotrophic or eutrophic ecosystems where nutrient availability and plankton biomass is
higher (Marbà et al. 2006; Krause-Jensen, Markager & Dalsgaard 2011). Plankton
detritus has lower C:N and C:P ratios than seagrass detritus (Duarte 1990), and this
difference in elemental composition could affect how quickly OM is mineralised, the
products of mineralization, and the physiological impacts of mineralization (Enriquez,
Duarte & Sand-Jensen 1993; Eyre & Ferguson 2002). Similarly, inputs of OM resulting
from anthropogenic activities will also have different chemical composition and lability,
yet the distinction between types of OM is rarely made. Inputs of OM into seagrass
meadows are often expected to cause detrimental effects based on previous experiments
that focus on highly degraded ecosystems (Delgado et al. 1999) or from experiments that
use unnatural OM forms that would not be found in natural seagrass sediments (Pérez et
al. 2007; Ruiz-Halpern, Macko & Fourqurean 2008), but the actual effects of OM
enrichment will depend largely on the dominant source of OM inputs. Characterising the
origins and chemical composition of OM inputs in seagrass sediments will be important
in predicting subsequent ecological effects.
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Seagrass ecosystems – a new focus on organic P
In this thesis, I show that the OM in seagrass sediments influences P cycling in seagrass
sediments, and that organic P may be an important contributor to the P requirements of
seagrasses growing in carbonate sediments that would normally be expected to be P
limited.
Models of P cycling and availability in seagrass ecosystems have focused on inorganic or
labile P pools in sediments and in the water column, primarily driven by an early focus
on the detrimental effects of inorganic nutrient enrichment in seagrass meadows (Short
1987). This focus has led to the paradigm of P limitation of seagrass growth in tropical,
carbonate sediments given the low SRP concentrations that are common in these
ecosystems, as well as the high adsorption of P onto calcareous sediment particles and N
being in excess due to nitrogen fixation (Short 1987; Romero et al. 2006). However, N
limitation has been found of seagrasses growing on carbonate sediments (Erftemeijer et
al. 1994; Udy, Dennison & Lee Long 1999), while P limitation has been found in seagrass
growing on sediments with low carbonate content (Pérez et al. 1991), showing factors
other than sediment type influence nutrient limitation.
When examining the availability and importance of sediment organic P in seagrass
meadows, I found it was important to couple traditional sequential extractions with
techniques such as 31P-NMR to examine the organic P fraction to a compound level. Only
a few studies had previously examined P fractionation in seagrass sediments (Jensen et
al. 1998; Koch, Benz & Rudnick 2001; Holmer, Carta & Andersen 2006; Reimer &
Huerta-Diaz 2010) and, in all of these experiments, the organic P fractions was
operationally defined based on extraction techniques. While providing valuable
information of P fractionation in sediments in general, this does not provide any
information on the potential turnover and availability of the organic P pool itself, given
that the variety of compounds that contribute to the organic P pool can vary greatly in
their availability and residence times in sediments (Turner, Frossard & Baldwin 2005;
Baldwin 2013). I found that the dominant organic P compound in Shark Bay sediments
were phosphate monoesters. Phosphate monoesters are not labile, and often accumulate
in soils and sediments with high amounts of plant litter (Turner et al. 2002), and their
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dominance in seagrass sediments is not surprising given that seagrass litter is often the
major component of OM in seagrass sediments (Cawley et al. 2012). Further research in
other seagrass ecosystems is required to determine if phosphate monoesters dominate the
organic P pool in all seagrass sediments, or if other organic P compounds may also form
a substantial proportion of the fraction. Again, the ubiquitous presence of alkaline
phosphatase – which could hydrolyse phosphate monoesters - in many different seagrass
ecosystems (Pérez & Romero 1993; Martínez-Crego, Romero & Alcoverro 2006; Koch,
Kletou & Tursi 2009) suggests that phosphate monoesters make up a substantial
proportion of organic P in these environments.
The composition of the organic P fraction in seagrass sediments will depend largely on
the types of organic inputs into meadows. Where other OM inputs such as planktonic- or
terrestrially-derived OM contribute substantially to the total sediment OM pool, the
composition of the sediment organic P fraction is likely to be different. For example,
sediment traps from a range of different coastal and open ocean ecosystems showed that
settling particulate matter consisted of phosphate monoesters, phosphate diesters, and
phosphonates (Paytan et al. 2003), while detrital inputs from adjacent coastal ecosystems
such as mangroves or wetlands would also have detectable concentrations of phosphate
diesters and phosphonates in the organic P pools (Cheesman, Turner & Ramesh Reddy
2012; Baldwin 2013). Further research examining the drivers of organic P speciation in
seagrass sediments would help determine the potential availability of this overlooked
nutrient pool.
Extreme events: an emerging threat to seagrass meadows
This thesis showed the drastic impacts that short-term, extreme climatic events like the
2010/11 marine heatwave can have on seagrass meadows. Though the effects of
increasing temperature on seagrass physiology and ecology have been widely studied
(Walker & Cambridge 1995; Campbell, McKenzie & Kerville 2006; Marbà & Duarte
2009), the impacts of increasing temperatures resulting from climate change has generally
been considered as a gradual, press disturbance where mean temperatures slowly
increase, giving seagrasses time to acclimate and adapt. However, extreme events that
lead to short lived periods of temperatures above mean levels are expected to increase in
frequency and intensity in the near future (Alexander & Arblaster 2009), and will present
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short-term pulse disturbances to seagrass ecosystems over time scales that limit
acclimation or adaptation. The capacity for seagrasses to persist through extreme events
will depend primarily on their upper thermal tolerance limits. Temperate seagrasses
growing near their upper thermal tolerance thresholds, like Amphibolis antarctica in
Shark Bay, will be most at risk to suffer negative effects following extreme events, and
may be replaced by smaller tropical seagrasses better suited for survival through high
temperature events. As such, seagrass ecosystems may respond to warming through
‘tropicalization’ similar to fish species, with transition zones from tropical to temperate
species gradually move polewards (Vergés et al. 2014). Impacts on seagrasses will be
particularly acute when extreme temperatures coincide with additional stressors. In this
thesis I showed that areas with reduced light availability had increased rates of dieback,
showing the capacity for multiple stressors to create patchy responses within an
ecosystem. Effective management of important seagrass ecosystems (especially at species
range limit) will require recognition of both pulse and press impacts of climate change,
and potentially with additive or synergistic stressors that would exacerbate loss. This may
require a shift of management strategies; away from direct strategies solely preventing
local loss driven by coastal development to strategies that also mitigate loss driven by
regional-global scale changes in climatology that acknowledge multiple stressors.
Shark Bay’s status as a World Heritage listed ecosystem is intrinsically linked to the
health and functioning of seagrass meadows within the Bay that could be affected by
future environmental changes (Kendrick et al. 2012). Clearly, extreme events like the
2010/11 heatwave have the potential to cause direct impacts to seagrasses, and therefore
key World Heritage values, within Shark Bay. Future extreme events could drive further
diebacks of temperate, meadow-forming seagrasses such as A. antarctica, which may be
replaced by smaller tropical seagrasses such as Halodule uninervis at sites that are
marginal for A. antarctica growth (Fraser et al. 2014). However, small tropical seagrasses
like H. uninervis do not offer the same ecosystem functions as A. antarctica, which is a
foundation species that is a niche habitat for species of high conservation priority in Shark
Bay (Fraser et al. 2014; Thomson et al. 2014). For example, the health of green turtles –
that use the seagrass meadows for foraging – decreased following the seagrass dieback of
Shark Bay (Thomson et al. 2014). Other species are also likely to have been impacted
following seagrass dieback in Shark Bay, but the full ecosystem effects are difficult to
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identify without long-term data sets to compare population changes before and after the
heatwave within Shark Bay. In addition, a shift towards a benthic community dominated
by tropical seagrasses would impact other ecosystem services such as sedimentation,
carbon sequestration, and nutrient cycling in Shark Bay (Walker, Kendrick & McComb
2006; Jahnert & Collins 2011; Fourqurean et al. 2012b). Clearly, any further loss of
seagrass in the Shark Bay World Heritage Area would have significant bottom-up impacts
on the ecologically and socially important functions that seagrasses contribute in Shark
Bay.
Microbial ecology of seagrass sediments
Microbial ecology and biogeochemistry are intimately linked in seagrass sediments, with
the functional composition of the microbial community largely determining the
availability of the organic P fraction and the decomposition of OM in sediments. In this
thesis I showed that microorganisms in seagrass sediments respond strongly to
environmental conditions (salinity and P availability), and shifts in microbial community
structure subsequently alter biogeochemistry and P availability in seagrass sediments.
However, microorganisms have a myriad of roles in plant rhizospheres that can influence
plant productivity; altering nutrient cycles, causing or increasing resistance to disease,
and regulating responses to potential stressors (Miki et al. 2010; Mangan et al. 2010; Van
Der Putten et al. 2013). The strength of such interactions would also have important
implications for species diversity within seagrass meadows; with strong positive
feedbacks likely to promote monospecific meadows and negative feedbacks promoting
meadows with higher species diversity (van der Heijden, Bardgett & van Straalen 2008;
Mangan et al. 2010; Van Der Putten et al. 2013). In seagrass sediments specifically,
microorganisms also mediate the availability of oxygen (Borum et al. 2005) and potential
toxins such as hydrogen sulfides in sediments that can influence seagrass productivity
and health (Pedersen, Binzer & Borum 2004). The interactions between seagrasses and
microorganisms underpin the functioning of seagrass meadows, yet these interactions
remain relatively understudied in seagrass ecosystems. As such, key questions regarding
microbial ecology in seagrass sediments remain unanswered. How do seagrasses interact
with microorganisms in sediments? What are the major drivers of microbial community
structure and function in seagrass rhizospheres? Are seagrass-associated microbial
communities species-specific, or do they share similarities among seagrass species? Are
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there key microbial groups that are critical to healthy seagrasses? I believe microbial
ecology is one of the most pressing areas of research in seagrass ecology, and emerging
techniques will allow us to address some of these key questions.
A rapid increase in our understanding of microbial ecology in seagrass ecosystems will
be facilitated by recent advances in molecular methods, which are increasing in power
and efficiency while decreasing in cost (Simon & Daniel 2011). In this thesis, I used
metagenomic sequencing to examine the taxonomic and functional structure of the entire
microbial community in seagrass sediments. Microbial community taxonomy and
function do not always change concurrently in marine ecosystems (Frias-Lopez et al.
2008). The functional diversity ultimately determines the microbial contribution to
ecosystem processes such as decomposition and nutrient cycling (Zak, Blackwood &
Waldrop 2006), yet most studies examining microorganisms associated with seagrasses
have focused on taxonomic measurements (Jensen, Kühl & Priemé 2007; Ikenaga et al.
2010; Green-García & Engel 2012). In the future, it will be more appropriate to examine
functional genes along with taxonomic structure in seagrass-associated microbial
communities when addressing the effects of microorganisms on ecosystem processes or
seagrass function. Complementary techniques such as metatranscriptomics and
metaproteomics can be used to examine changes in gene and protein expression
respectively (Wilmes & Bond 2006; Gilbert et al. 2008), and used alongside molecular
methods like metagenomics to show real time microbial community responses to changes
in physical conditions. In situ hybridisation methods such as fluorescent in-situ
hybridization (FISH) and high resolution secondary ion mass spectrometry (nano-SIMS)
could identify specific microbial groups and their locations within seagrass tissues and in the case of nano-SIMS - allow us to determine their importance for nutrient transfer
into seagrasses using stable isotope labelling (Amann, Fuchs & Behrens 2001; Behrens,
Kappler & Obst 2012). These techniques could be combined with manipulative
experiments under controlled conditions to explicitly investigate the importance of
microorganisms to seagrass physiology. Plant-soil feedback experiments have become
increasingly common in terrestrial plant ecology, and involve experimentally altering soil
microbial communities (e.g. sterilizing sediments or preconditioning with another species
prior to planting) and testing the effects on plant health (Van Der Putten et al. 2013).
These techniques will allow us to move beyond simply cataloguing what microbial
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species are present in seagrass meadows, and instead focus on the ecological functions
they perform.
A new assessment of the biogeochemistry of Shark Bay
This thesis has revealed new insights in to the dependence of biogeochemical processes
and microbial functioning of Shark Bay that demonstrate a previously unappreciated role
of seagrass OM in maintaining P availability. Previously, Shark Bay has been widely
considered a notable example of an oligotrophic marine embayment, where primary
productivity is strongly P limited, especially at high salinity sites (Smith & Atkinson
1983; Smith 1984a; b). This thesis shows a clear shift in the predominance of different
sedimentary P fractions (Chapter 4) and microbial community function (Chapter 5) along
the salinity gradient of Shark Bay that changes the availability of sedimentary P, and may
explain why seagrasses do not always show signs of increasing P limitation across the
salinity gradient (Fraser et al. 2012). Patterns in sediment P speciation, microbial P
metabolism gene abundance, and enzyme expression suggest microbial communities
were increasingly targeting the organic P fraction in high salinity areas.
This thesis also further shows the capacity of seagrasses to act as ecosystem engineers by
modifying their surrounding environment. Seagrasses are generally considered ecosystem
engineers through their role modifying the physical structure of the environment (e.g. Bos
et al. 2007), as they do in Shark Bay by creating shallow banks through increased
sedimentation (Kendrick et al. 2012). However, this thesis demonstrates that the
seagrasses of Shark Bay also contribute to biogeochemical gradients in sediment OM and
P availability, as well as salinity, creating a unique environment in southern, hypersaline
reaches where they are able to dominate as primary producers in spite of extreme
conditions of oligotrophy and hypersalinity. Sediment microorganisms are likely an
important part of seagrasses creating and surviving within such extreme environments;
containing novel genes and potentially forming close associations with seagrasses that
enable nutrient demands to be met from organic sources.
If recycling of nutrients from seagrass-derived OM is important in supporting primary
productivity in Shark Bay, further loss of seagrasses would impact biogeochemical
cycling in Shark Bay. Seagrasses provide microorganisms with a unique environment by
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Phosphorus cycling in seagrass sediments
creating a gradient in oxygen around roots and through enriching sediments with seagrass
trapped and derived OM (Welsh 2000; Garcia-Martínez et al. 2009). Interactions between
plants and microorganisms also become increasingly important for plant nutrition under
oligotrophic conditions, while functional redundancy is decreased for important and often
highly specialized functions (van der Heijden et al. 2008). Consequently, Shark Bay may
be highly sensitive to reductions in microbial diversity arising from shifts in
environmental conditions. Predicted increases in the magnitude and frequency of extreme
events like the 2010/11 heatwave and floods in Shark Bay would directly influence
microbial communities by changing environmental conditions, and resultant significant
seagrass dieback would also indirectly impact seagrass-associated microbial communities
through a loss of habitat (Jensen et al. 2007; Williams et al. 2009). Extreme climatic
events leading to increased sea temperatures and flooding will increase in Shark Bay
(Alexander & Arblaster 2009; Ummenhofer et al. 2015), while global ocean temperatures
are expected to increase 0.2°C per decade under most projections (IPCC 2007). In
addition, associated changes to ocean circulation (Hobday & Lough 2011) and hydrology
(Ummenhofer et al. 2015) could influence salinity or nutrient availability that would also
likely impact seagrass and microbial communities. The subsequent effects on
biogeochemical processes, especially of phosphorus and carbon in Shark Bay are hard to
predict. For example, increased temperatures could increase seagrass-derived OM in
sediments by causing seagrass dieback, but would also increase mineralization rates of
existing sediment OM.
Shark Bay has limited pressures from surrounding anthropogenic activities, which has
contributed to it remaining a relatively pristine ecosystem (Kendrick et al. 2012).
However, Shark Bay would be particularly vulnerable to future nutrient inputs. Currently,
Shark Bay has been largely protected from eutrophication issues faced my many other
seagrass communities around the world owing to its extreme isolation (Burkholder,
Tomasko & Touchette 2007). However, even small nutrient additions can lead to changes
in benthic community structure away from temperate, structurally-important seagrasses
to tropical seagrasses, macroalgae, or even unvegetated sediments (Armitage et al. 2005;
Armitage, Frankovich & Fourqurean 2011). Increases in nutrients in readily available
inorganic forms like nitrates or phosphates could rapidly alter benthic biogeochemistry
and ecology in surrounding meadows, especially since biogeochemical cycling within
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Shark Bay seems dominated by internal recycling processes mediated by microbial
activity. As such, any future development that could increase nutrient inputs (e.g.
population growth, increased tourism) could have significant negative impacts on the
seagrass meadows of Shark Bay, and must be managed accordingly. Nutrient addition
experiments would show how benthic communities would respond to potential changes
in nutrient inputs, allowing predictions of ecosystem responses that would inform the
future management of Shark Bay.
Conclusion
Nutrient cycling in seagrass sediments is a major driver of seagrass ecology, yet the
contribution of nutrients from organic pools have largely been ignored, in spite of
seagrass sediments often being enriched in OM. This thesis showed that OM in seagrass
sediments has significant effects on biogeochemical processes and microbial ecology in
the oligotrophic seagrass meadows of Shark Bay, and has the potential to drive changes
in seagrass productivity. Even seagrass-derived OM – previously considered as refractory
in seagrass sediments –greatly alters belowground processes, and will likely become more
prevalent in key seagrass ecosystems like Shark Bay with increases in extreme climatic
events. Interactions between sediment OM, microbial communities, and seagrasses
should be further elucidated to better understand the belowground drivers of seagrass
ecology which have been often overlooked in favour of more visible changes in water
quality. Given that belowground interactions have the potential to alter nutrient budgets,
primary productivity, and even seagrass diversity, this should be a research priority. Plant
ecology, sediment biogeochemistry, and microbial ecology are intrinsically linked to one
another by a series of complex interactions that together determine key ecosystem
functions like primary productivity, and must be examined concurrently to understand
belowground drivers of coastal ecosystem ecology. Quantifying the link between
microbially derived recycled organic P and uptake of P by seagrasses remains the missing
link in this thesis given the methodological difficulties with extracting δ18OPO4 from
seagrasses and sediments, but deserves immediate further investigation.
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Amann, R., Fuchs, B.M. & Behrens, S. (2001) The identification of microorganisms by
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Phosphorus cycling in seagrass sediments
Appendices
143
144
Depth
(m)
1.8
1.6
1.9
1.3
1.5
1.6
1.5
2.3
1.8
Site
1
2
3
4
5
6
7
8
9
25.2
25.3
26.1
26.0
24.4
25.0
26.4
26.3
25.7
Temp
(°C)
BD
BD
BD
0.05
0.06
0.08
0.06
0.1
Water SRP
(µM)
0.1
51.2
53.2
46.8
42.7
43.6
42.3
40.6
39.7
39.2
Salinity
(‰)
2.1
2.8
4.3
4.1
2.0
4.3
1.5
2.7
1.7
Organic
matter
(% d.w)
239.0
702.3
493.3
131.0
66.8
102.2
26.5
31.0
123.9
Alkaline
phosphatase
(µmol-1 hr-1)
79.3
230.0
202.9
71.0
39.3
43.3
17.1
23.9
40.4
Acid
phosphatase
(µmol-1 hr-1)
3.1
11.2
16.7
17.4
4.7
8.5
3.4
4.9
4.7
Beta
glucosidase
(µmol-1 hr-1)
54.7
30.5
66.4
52.3
15.5
7.3
8.1
16.0
35.0
Microbial
P biomass
(µg g-1)
0.06
0.07
0.05
0.07
0.06
0.05
0.05
0.05
0.06
Seagrass
P (% d.w)
Appendix 1: Summary of physical and biogeochemical data at each site. Depth, temperature, and salinity all averages over 5 CTD casts. Sediment organic matter,
enzyme expressions, microbial P biomass, and seagrass P content are averages (n=3 per site). SRP = Soluble reactive phosphorus. BD = Below detection
Fraser (2016)
Phosphorus cycling in seagrass sediments
Appendix 2: Post-hoc analysis (Games-Howell) of differences in abundance of level 1
functions annotated using the SEED database between high salinity sites (>46‰, white)
and lower salinity sites (<46‰, dark grey). Level 1 functions that were significantly
different between high and lower salinity sites are denoted by *, with corrected P-values
using Storey’s false discovery rate approach (P<0.05, see Chapter 5 methods for full
description).
145
146
4661589.3
4661588.3
4661587.3
4661598.3
4661599.3
4661602.3
4661601.3
4661597.3
4661596.3
4661603.3
4661593.3
4661604.3
4661591.3
4661590.3
4661594.3
4661600.3
4661595.3
4661592.3
MG-RAST ID
9
9
9
7
7
7
6
6
6
5
5
5
3
3
3
1
1
Site
1
1,207,337,003
1,197,907,978
1,167,185,047
1,209,795,764
1,186,351,414
1,206,253,769
1,202,976,906
1,224,919,206
1,184,808,271
1,214,421,809
1,192,194,650
1,160,096,662
1,197,070,339
1,125,452,740
1,167,798,847
1,233,436,148
1,214,340,697
1,466,828,783
Base pair count
3,841,773
3,757,584
3,599,493
3,786,067
3,678,190
3,777,840
3,755,410
3,854,469
3,672,117
3,754,913
3,714,494
3,634,852
3,753,386
3,495,614
3,528,382
3,836,447
3,776,363
4,493,000
Sequence count
314
318
324
319
322
319
320
317
322
323
320
319
318
321
330
321
321
326
Average
length
sequence
33.4
32.6
30
31
31
32.2
33
30.5
30.1
33.7
31.7
25.5
30.5
25.5
17.7
32.6
36
38.2
Annotated
protein (%)
1,144,768
1,175,768
1,013,308
1,028,746
1,084,656
1,124,935
1,101,648
1,040,510
1,029,926
1,207,849
1,149,229
853,589
1,006,349
845,917
301,420
1,281,710
1,333,736
1,664,043
Number
subsystem
proteins
37
39
34
33
35
36
34
33
33
38
38
31
35
33
12.3
44.2
43
47
Predicted
Subsystems proteins
(%)
727
743
768
724
740
719
743
741
753
725
748
701
706
757
758
696
718
671
α-diversity
(species)
Appendix 3: Summary of metagenomic sequencing results for Shark Bay sediments. Number of base pairs, sequencing reads, annotated proteins and % predicted using
the SEED subsystem database after quality control on the MG-RAST pipeline.
Fraser (2016)
Stromatolite
Sponge
Salt marsh
Rainforest soil
Mangroves
Grassland
Freshwater
Estuary
Desert
Deep Sea
Coral
Biome Type
Seagrass
4604137.3, 4604139.3, 4604141.3
4530252.3, 4530290.3, 4530370.3
4520021.3, 4520022.3, 4520023.3
4497403.3, 4497404.3
4523017.3 – 4523020.3
4539064.3, 4541651.3
4465820.3, 4465821.3, 4465822.3
4440948.3, 4441020.3 - 4441022.3,
4487294.3, 4487295.3
(i) 4477805.3,
(ii) 4477872.3-4477873.3
Shark Bay, Western Australia
Botany Bay, Australia
Plum Island, USA
Amazon Forest, Brazil
Red Sea, Saudi Arabia
Cedar Creek, USA
Colorado River, USA
Coorong Lagoon, South Australia
Offshore sediment, South China Sea
(i) Mojhave Desert, California
(ii) Chihuahuan Desert, New Mexico
Magnetic Island, Australia
Shark Bay, Western Australia
4661587.3 – 4661604.3
4445755.3 – 4445756.3
Location
MG-RAST IDs
Ilumina
454
Ilumina
Ilumina
454
Ilumina
454
Ilumina
454
Ilumina
454
Sequencin
g method
Ilumina
3
3
3
2
4
2
3
4
3
2
2
18
Samples
Appendix 4: Summary of metagenomes used to compare microbial communities in seagrass rhizosphere sediments to other ecosystems. All
metagenomes are publicly available on the MG-RAST server.
Phosphorus cycling in seagrass sediments
147