Changes in the species and functional diversity of birds in tropical secondary forests

It’s been a while since I’ve posted anything on here and I’ve been feeling a bit bad about it. My excuses (and I think they’re good ones) are that I’ve been busy buying and fixing up a house and preparing for impending fatherhood (my partner is pregnant with twins and is due later on this month). I’ll try to post on here a bit more regularly over the next year, but I imagine that I’ll be quite busy*.

Anyway, a few months ago my master’s student Catherine Sayer had a paper we worked on during my PhD (and for a good while afterwards) published in Biological Conservation. She did an awesome job – I’m really happy to get this paper out there and feel like it produced some novel and interesting results.

We started the work with the idea of looking at how various aspects of bird biodiversity change during recovery of secondary tropical forests, as a kind-of companion piece to my paper on changes in plant biodiversity and biomass in secondary forests. Although people had previously looked at how bird biodiversity changes during secondary forest succession these studies almost exclusively looked used species richness as a metric, which as I’ve said previously isn’t necessarily that useful for telling us what is going on in an ecological community or for conservation. As a result, we decided to look at how species diversity and functional diversity – diversity defined by what species do in an ecosystem rather than their taxonomy – changed during recovery of tropical forests from agricultural clearance and how this compared to diversity in nearby primary forests.

After a systematic search, Catherine found 44 studies from which she could access raw data on species counts in secondary forests and nearby primary forests. As you can see from the map below, these were largely in South and Central America, with a number of other studies in Asia and relatively few in Africa. These secondary forests ranged from ones only a few years old to ones that were nearly a century old. We combined this with a great, global data set on bird traits to calculate a number of metrics of functional diversity.

Our most striking results were that, relative to primary forests, the species richness of forest specialist species in secondary forests increased over time since last disturbance and that the standardised effect size of functional diversity declined over time.

Broadly speaking this suggests that forest specialists increase in richness over time, that secondary forests resemble more mature forests as they age and that they can potentially support populations of specialists that may be threatened by deforestation. Regarding changes in the standardised effect size of functional diversity, what this means is that functional diversity was lower than expected relative to the number of species in bird communities for older secondary forests. This means that there were more species that were functionally similar in older forests, meaning that these species were likely to be carrying out similar processes. This is important because it means that the effects of population declines of one species on ecosystem process could be buffered by another similar species. As a result, ecosystem processes in secondary forests, and ultimately the ecosystem services they provide, may become more resilient as these forests age.

Our assertion that ecosystem processes in secondary and primary forests might be similar was also supported by our finding that there was relatively little difference between secondary and primary tropical forests when comparing other metrics of functional diversity.

These metrics did not appear to respond as strongly to forest age as the metrics mentioned earlier. However, even though we didn’t find a relationship with forest age I’d be surprised if they didn’t change during succession – I just think that there was too much noise in our data to pick up a signal. Factors such as the length and intensity of previous disturbances, connectivity, and size of forest patches are all likely to be important in determining how secondary forests change over time but unfortunately, most studies didn’t collect this data and so we couldn’t address these issues in our analysis.

From the results of our paper, it seems that primary forest is particularly important to conserve forest specialists, and if this is a priority it is clear that primary forests must be protected (as previous papers have also suggested). However, it also appears that secondary forests have the potential to provide similar ecosystem processes, and possibly services when compared to primary forests. Even if this is the case, older secondary forests are more likely to be resilient to disturbances than young secondary forest and so any regrowth should be protected wherever possible, particularly in areas where little primary forest remains.

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*This is an understatement, I’m led to believe that I should expect around three hours sleep a night for the first few months…

Is it useful to categorise disturbances as pulses or presses?

In the 1980’s Edward Bender coined the terms ‘pulse’ and ‘press’ to describe different types of disturbance to ecological communities. These terms were originally intended to be used to classify experimental perturbations. However, in the 30 years since they were first used this division between pulse and press has become the dominant way of thinking about perturbations in real-world ecosystems. Bender defined pulse disturbances as perturbations from which the community is subsequently allowed to respond. Now pulse disturbances are seen as any relatively short-term, easily delineated disturbance.  Press disturbances, on the other hand, were defined by Bender as those in which ‘densities of perturbed species are altered and maintained at predetermined levels by adding or removing individuals as needed.’ This has since been altered with press disturbances defined as those that continue at a similar intensity following their initial occurrence*. Common examples of pulse disturbances are fire, drought, or selective logging, while press disturbances may represent conversion to agricultural use. Importantly, deciding whether a disturbance is classified as a pulse or a press comes down to whether a researcher considers its duration to be short-term or not.

As a result, this definition  of what category a disturbance falls into is inherently subjective. Pulse disturbances are often events that happen in a single year, but is a drought that lasts for three years still a pulse? How about conversion to agriculture that is abandoned after a decade, is that still a press disturbance? What if it is abandoned after a year, is it a pulse disturbance now? All of this results from an imprecise definition, which can lead to serious misunderstandings in ecology. However, this is not an issue that can be solved a better definition. Rather, I think that dividing disturbances into categories is unhelpful and the categories pulse and press do not communicate all the details required to compare disturbances.

As I’ve written before, I’m not a big fan of categorising things in ecology when it can be avoided. My opposition comes from the fact that I think it removes a lot of the nuance regarding what is going on in ecological systems (Brian McGill has written a great post on this issue). Comparing the impact of pulse and press disturbances on biodiversity may be superficially interesting, but it’s much more interesting to look at how the duration of a disturbance influences biodiversity change and any subsequent recovery . Ideally, when investigating the influence of disturbances it would be useful to know how (i) long a disturbance lasted (or how long since it started in the case of ongoing disturbances) and (ii) when a disturbance has ceased, how much time has passed since the last disturbance. In a perfect world, we would also like to know the intensity of the disturbance, but we can usually only find this out by analysing what happens to an ecological community in response to the disturbance.

Another reason not to classify disturbances as pulse or press is that this doesn’t take into account any of the ecological differences in taxonomic groups you may be investigating. For example, if a drought occurs in a temperate forest in a single year it will be experienced as a temporary change for a tree which may survive for hundreds of years. However, this drought may represent the entire lifespan of an invertebrate species. So even though the disturbance length for the two species may be the same, the drought may account for <1% of the lifespan of the tree while accounting for ~100% of the invertebrate’s lifespan. Therefore, the same disturbance may have very different effects on different species groups. If the disturbance inhibits reproduction this would result in steeper declines for species with a naturally shorter lifespan**.

The same issues arise when explaining the effects of longer lasting change, such as agricultural conversion. All else being equal declines would be expected to be more pronounced in species with shorter life spans, and recoveries in their populations would likely be faster. This has already been recognised by work on extinction debt, which suggests that species with short generation times are less likely to show lags between disturbance and population decline than longer-lived species.

So, although dividing between pulse and press disturbances may be superficially appealing, reporting the duration of disturbances is likely to allow greater potential for generalisation. Combining this data with information on species traits such as mass/longevity will then help to answer the question of how the effect of disturbance duration differs amongst species and which species might be most vulnerable to anthropogenic change. This approach could also help to improve our understanding of d recovery following disturbances, a process which will likely become more widespread as humans continue to move from the countryside to cities.

* The term ramp disturbance was also introduced by Lake in 2000 to describe disturbances that increase in intensity over time, such as climate change, but I didn’t want to get into that here.

**Some of what I’ve written here is a result of some back-and-forth I had with Martin Jung and Dale Nimmo on Twitter, you can see a summary of that conversation here.

How long does it take for logging roads to recover from clearance?

Roads are generally terrible for biodiversity. They fragment habitat, can increase habitat loss and hunting as a result of increased access, and cause direct loss of biodiversity as a result of collisions. However, not all roads are the same. Some are massive, permanent structures, while others are temporary, dirt tracks that may seemingly disappear once they fall into disuse.

One example of ephemeral roads is those that logging companies construct in tropical forests to provide access and transport of logs. There has long been concern that these roads can increase the risk of fires occurring, as well as increasing access for hunting, and other forms of forest exploitation. However, in a recent(ish) paper* has shown that some of the negative effects of logging roads are relatively transient.

In the paper, Fritz Kleinschroth and colleagues showed that in Central African forests, after 30 years of recovery logging roads had similar canopy cover, species diversity, and leaf litter to logged forests nearby* . However, the amount of carbon stored in the form of biomass lagged behind and was only 6% of that seen for logged forests after 30 years of recovery. At this rate, biomass recovery would take more than 300 years.This incredibly slow recovery at first appears puzzling, given that secondary forests, which have had almost all their trees cut down in the past and turned into agricultural fields, tend to take between 60-100 years to recover biomass to pre-disturbance levels (see here for a blog post and here for a recent paper on this). However, the probable cause of this delayed recovery is the compaction of soils on the roads by heavy vehicles which reduces seed germination and root growth***. Taken together the authors suggest these results indicate that logging roads have the potential to act as areas in which timber species could regenerate and that they may become inaccessible to hunters within 10 years.

So how does this study compare to similar ones carried out previously? Firstly, this study is a little different because it is one of the few that used chronosequences to assess recovery, and so the only study I know of which can assess longer-term dynamics on logging roads. However, other studies present a similar picture for the recovery of biomass and forest structure – forest canopy cover recovers relatively quickly (see here and here), but biomass and basal area lag behind (see here and here). The major difference between this study and previous ones is that it presents a more optimistic outlook of biodiversity. Previous studies have estimated that species richness may be 50-95% lower on abandoned logging roads when compared to logged forests (see here, here, and here). As such, the relatively fast recovery of species richness shown by Kleinschroth and colleagues appears to be outside of the norm, and further similar studies will be needed to see whether the pattern of recovery shown in this paper is an outlier. So we can’t really give a solid answer to the question posed in the title of this post – sorry about that.

While results vary from study to study it is obvious that more efforts need to be made to reduce the number of logging roads, their initial impacts to forests, and to help them recover once they are abandoned. In order to reduce the number of roads, reduced-impact logging could be used. This method, which I have been accused of disliking in the past, seems to be very successful in reducing the number of roads in logging concessions where it has been used (see here for an example of this). This is done by producing a plan for road construction prior to any trees being cut, rather than the ad-hoc approach often taken in conventional logging. Reducing the impact of roads could be done by limiting their width. Finally, improving recovery could be done by planting seedlings/saplings on former logging roads, as well as reducing access to roads.

One suggestion that the authors made in their paper that I really like is to re-use logging roads when forests are re-logged. Given that logging typically occurs every 30 years, this would allow some time for recovery of biodiversity on the roads but clearing them would reduce the damage caused by their construction spreading to other areas of the forest.

*I admit it, I’ve been terrible at keeping up with my reading recently.

**John Healy and Fritz have written a nice summary of their paper on the website “The Conversation” which is well worth a read.

***Anecdotal, I know, but I have seen similar things on restored salt marshes in the UK where diggers have been used to breach sea walls. At least for the ones I remember, this resulted in reduced vegetation cover.

 

Are we in danger of underestimating biodiversity loss?

Almost all ecological research of human impacts on biodiversity looks at changes after they have happened. To do this, researchers usually compare a site where some kind of disturbance has happened to a nearby undisturbed site. This method is called space-for-time substitution. The assumption of this approach is that the only thing that differs between sites is this disturbance. However, this assumption is often incorrect. Sites may have had very different biodiversity before any disturbances, which can lead to under- or over-estimations of biodiversity changes as a result of human impacts. One result of this is that we aren’t really sure how tropical logging alters the composition of ecological communities. These problems are likely to be particularly acute when habitat fragmentation limits dispersal to some sites.

Up until recently there had been little work comparing how the results from space-for-time methods compare to methods that compare sites before and after disturbances. However, last week an elegantly designed study was published in the Journal of Applied Ecology which aimed to examine just this in the context of logging in Brazil. The paper aimed to compare space-for-time methods to before-after-control-impact (BACI) methods. Critically BACI studies measure biodiversity at sites at least once before the disturbance of interest takes place. Researchers then return to sites and remeasure them after the disturbance. Importantly both sites impacted by the disturbance and control sites are surveyed on both occasions. Using this method allows researchers to disentangle the effects of disturbances and any differences between sites prior to disturbance – a key advantage over space-for-time methods.

The paper by Filipe França and colleagues examined the differences in results obtained for space-for-time and BACI methods when looking at changes in dung beetle biodiversity in tropical logged forests in Brazil. To do this they surveyed 34 locations in a logging concession, 29 of which were subsequently logged at a variety of intensities. The intensity of logging (the number of trees/volume of wood removed per hectare) is a very important determinant of the impact of logging on biodiversity and carbon (see previous posts on this here and here). They then went back and re-surveyed these locations one year later. From the data collected, they calculated changes in species richness, community composition and total biomass of dung beetles.

Figure 1 – Differences between before-after-control-impact (BACI) approach and space-for-time substitution (SFT) for changes in dung beetle species richness, community composition, and biomass. For more details see Franca et al. (2016).

When comparing space-for-time and BACI the paper found that BACI characterised changes in biodiversity significantly better than space-for-time methods. Critically, space-for-time methods underestimated the relationship between logging intensity and biodiversity losses, with changes in species richness twice as severe as estimated by space-for-time (see Figure 1). BACI methods also consistently provided higher explanatory power and steeper slopes between logging intensity and biodiversity loss.

So what does this mean for how we do applied ecology? I think it is clear that we need to employ BACI methods more often in the future. However, BACI comes with logistical and financial constraints. Firstly, it is virtually impossible to predict where disturbances are going to happen before they occur. As a result, Franca and colleagues think that if we want to carry out more BACI research in the future, we need to develop closer ties with practitioners. This will involve building relationships with logging and oil palm companies, as well as agricultural businesses and property developers. This may make some researchers uncomfortable, but we need to do this if we are to provide robust evidence for decision makers. Secondly, BACI studies take longer to carry out, so we need to convince those that hold the purse strings that they are worth investing in.

BACI is clearly something we should be using more often but does this mean that space-for-time approach is useless? Should we even be using space-for-time methods at all? I’m not being hyperbolic just to get some attention- some have argued that we should stop using chronosequences altogether because ecological succession is unpredictable. After momentarily going into a bit a crisis about this when I read some papers on succession last year, I have come to a slightly different conclusion. Space-for-time substitution sometimes predicts temporal changes well, but sometimes it doesn’t. What we need is to work out when the use of space-for-time approaches are acceptable, and when it would be better to use temporal methods. Reviews have highlighted that as ecosystems increase in complexity space-for-time methods become less useful for monitoring changes in biodiversity. For example, large local species pools mean that post-disturbance colonisation may be very variable between sites. This problem is  compounded in fragmented landscapes where there are barriers to dispersal of seeds and animals. Every additional layer of complexity makes post-disturbance dynamics more and more difficult to predict. Ultimately, the best way to address this problem is through some kind of synthesis.

Working out when space-for-time approaches are useful and when they are not is not something we are going solve overnight. Before we can review the evidence, we need some evidence in the first place.  This is part of the reason why papers like the one by França and colleagues that I’ve discussed here are vitally important. So next time you think about designing a study see if you can assess how the results from temporal methods compare those from  space-for-time methods. The results might just take you by surprise.

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Filipe França & Hannah Griffiths have written a great post on the Journal of Applied Ecology blog going into more detail about the implications of their study. I strongly recommend you give it a look.

Local species richness may be declining after all

Recently two papers seemed to turn what we thought we knew about changes in biodiversity on their head. These papers by Vellend et al. and Dornelas et al. collated data from multiple sources and suggested that species richness at local scales is not currently declining. This was counter-intuitive because we all know that species are going extinct at unprecedented rates. However, it is possible that the introduction of non-native species and recovery of previously cultivated areas may offset extinctions leading to relatively little net change in local species richness.

This week a paper has been published that calls these findings into question. The paper by Andy Gonzalez and colleagues published in the journal Ecology, suggests that there are three major flaws with the analyses. These flaws mean that the answer to the question ‘Is local-scale species richness declining?’ currently remains unanswered and is unanswerable.

The papers of Vellend et al. and Dornelas et al. were meta-analyses of previously published papers. One issue with meta-analysis is that it is very prone to bias. Like any study if the samples (in this case ecological studies) are not representative of the population (in this case locations around the globe) then any results will be flawed. To test the representativeness of the datasets used by Vellend and Dornelas Gonzalez et al. examined how well they represented biodiversity and threats to biodiversity. This analysis (see below) showed that the papers were not representative of biodiversity or the threats faced by biodiversity (though curiously, the analysis of Dornelas et al. showed an overrepresentation of areas highly impacted by human impacts).

Figure 1 – Spatial bias of the Vellend et al. (2013) and Dornelas et al. (2014) data syntheses. For more information see the paper by Gonzalez et al. (2016).

The paper also suggests that using short time series can underestimate losses. By analysing the effect of study duration and changes in species richness (see below) Gonzalez et al. claim that increases in study duration were correlated with a decline in species richness. This supports previous theory which suggests that there is often a time lag between disturbance events and species extinctions – termed ‘extinction debt.’ However, I’d be intrigued to see the results of removing the studies with the longest duration from this analysis since the authors admit that the analysis is sensitive to their inclusion. I’ve seen recent similar work that suggests the same kind of relationship might be seen for studies monitoring individual animal populations.

Figure 2 – The effect of study duration on apparent changes in species richness.

Thirdly, Gonzalez et al. assert that including studies in which ecosystems were recovering from disturbance (e.g. regrowth on former agricultural fields) without taking into account historical losses that occurred during or after the disturbance biases estimates of change. The paper by Vellend et al. in particular combined studies of the immediate response of biodiversity to disturbances such as fire and grazing along with studies of recovery from the very same disturbances. Gonzalez et al. show that once studies of systems that were recovering are removed from Vellend et al’s analysis there is a negative trend in species richness changes.

The biases prevalent in the Vellend and Dornelas papers lead to Gonzalez et al. to suggest that the papers cannot conclude what the net changes in local species richness are at a global scale. However, they note that the results of Dornelas and Vellend are in sharp contrast to other syntheses of biodiversity changes which used reference undisturbed such as those by Newbold et al. and Murphy and Romanuk which reported average losses of species richness of 14 and 18% respectively.

In their conclusion Gonzalez et al. suggest that though meta-analysis is a powerful tool, it needs to be used with great care. Or to put it another way, with great power comes great responsibility. As someone who regularly uses meta-analysis to form generalisations about how nature works I completely agree with this statement. Traditionally scientists have used funnel plots (graphs with study sample size on the y-axis and effect size on the x-axis) to identify biases in their analyses. I’ve always been skeptical of this approach, especially in ecology where there is always a large amount of variation between sites. In the future syntheses would do well to follow the advice of Gonzalez et al. and really interrogate the data they are using to find any taxonomic, geographic, climatic or any other biases that might limit their ability to generalise. I know it’s something I’ll be taking more seriously in the future.

Gonzalez et al. also point out that most ecological research is carried out in Europe and North America. If we want to monitor biodivesity we need to increase efforts in biodiverse tropical regions, as well as boreal forests, tundra and deserts. We need to identify where these gaps need filling most and then relevant organisations need to prioritise efforts to carry out monitoring. I am positive that this can be achieved, but it will cost a lot money, needs to be highlighted as a priority and will ned a lot of political good will. Even with this effort some of the gaps in biodiverse regions, such as the Democratic Republic of Congo, will be extremely difficult to fill due to ongoing armed conflict

My take-home message from this paper is that we need to be more careful about how we do synthesis. However, I also think that species richness isn’t the only metric that we should focus on when talking about biodiversity change. Studies have shown that measures of the traits of species present in a community are generally more useful for predicting changes in ecosystem function than just using species richness. Species richness is the iconic measure of biodiversity, but it probably isn’t the best. Ecologists should view species richness in the same way as doctors view a thermometer – it’s a useful tool but you still need to be able to monitor blood pressure, take biopsies and listen to a patient’s lungs before you diagnose them*.

 

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*Thanks to Falko Bushke whose analogy I stole from a comment he made on my blog post here.

 

Second growth:The promise of tropical forest regeneration in an age of deforestation

Anyone who knows anything about secondary forests will have come across the work of Robin Chazdon. She has inspired at least one forest ecologist, me, that forests recovering from major disturbances are a subject worthy of study. I’m sure she has done the same for many others out there. So, coming towards the end of my PhD I was excited to see a book that she had written summarising the topic was due to be released and using the last of my NERC funds I bought it. And then I moved house to Spain, where the book sat untouched and unloved in a box for the next year. After I came back to the UK last year, I found the book again and decided I should stop putting off reading it. I read it on trains, buses, on my sofa and occasionally in bed. I once fell asleep reading this book, though admittedly that was on the way back from the BES annual meeting  in Edinburgh, and the gin from the previous night was probably the cause of my sleepiness rather than any bad writing.

The first thing to say is that this book is extremely comprehensive. Though it is not particularly lengthy, running to 316 pages of text, it covers a huge range of topics relating to forest regeneration from traditional knowledge and prehistoric forest transformations by humans to recovery pathways from fire, landslides, volcanic eruptions, logging, and agricultural use. There are also numerous sections on community assembly, functional traits, ecosystem function, and animal and plant interactions. The last section concentrates on reforestation and restoration of degraded forests, making a passionate plea for degraded forests to not be considered as wasteland.

For me the most fascinating parts of the book were those that covered traditional knowledge of forest regeneration and the history of human cultures in tropical forests – both subjects I knew practically nothing about before this book. I was captivated to read that the dayak people of Borneo have five words to define different stages of forest recovery – kurat uraq (1-3 year old scrub that forms after abandonment), kurat tuha (trees > 5 cm in diameter and 5-6 metres in height), kurat batang muda (trees 10-15 cm in diameter), kurat batang tuha (closed canopy secondary forest) and hutan bengkar (primary forest). As Chazdon points out this knowledge shows a striking resemblance to that of forest ecologists. Similarly, Mayan cultures in Central America and Soliga people in the Western Ghats have developed a subtle knowledge of the stages of forest succession. I have always been a bit skeptical of integrating traditional knowledge into ecological science, but this book convinced me that there could be some value to it.

Chazdon masterfully weaves together anthropology, archaeology and ecology in the discussion of prehistorical impacts of humans on tropical forests. She cites evidence of earthworks called geoglyphs similar to the Nazca lines found in the state of Acre in Brazil, swidden agriculture 20,000 years ago in Papua New Guinea and human populations in Central America to dispel the view that any forest is truly untouched. There are probably legacies of human use in most forests, we just can’t identify them. Based on this she, perhaps controversially, critiques recent work suggesting that mature tropical forest biomass density is increasing as a result of atmospheric carbon dioxide. Chazdon’s view is that this increase could well be as a result of recovery from unseen disturbances that happened generations ago.

The section on community turnover during succession is also excellent, with a detailed analysis of the characteristics of short- and long-lived pioneer and shade tolerant, late successional species. At points Chazdon playfully conjures up text resembling Shakespeare’s  “All the world’s a stage” monologue: “The term successional stage is apt. Successional pathways can be viewed as an improvisational drama in several acts, with each act featuring a different set of actors. Some actors perform throughout the drama, but others have cameo appearances of only one act. Although each act sets the stage for the next, forest regeneration has no director and only a roughly sketched script creating a high degree of spontaneity, randomness and uncertainty.” These are amongst my favourite parts of the book, with metaphor mixing with a solid science to help things stick in your mind that might otherwise be easily forgotten.

If I have any criticism of the book, it is that it’s a bit repetitive. This is probably because Chazdon sees succession as ‘an improvisational drama in several acts’ and so the book relies on case studies, rather than synthesising current knowledge to form generalities. However, I think that the repetition helps if you just want to dip in and out of chapters – I don’t think it is necessarily written to be read cover to cover like I did.

That aside if you are interested in the dynamics of forests in any way this book is essential reading. There is no better summary of current thinking on tropical forest succession out there.

“Like walking through an open cemetery”

“I have been working in human-modified tropical forests for the past 14 years, but seeing these fires first hand was devastating,” wrote Erika responding to one of my questions “The smell of wet soil was gone and I could only smell smoke…even the usual cacophony of forest sounds disappeared…it was like walking through an open cemetery.”

Erika de Berengeur Cesar, an up and coming Brazilian forest researcher, works at Lancaster University. For last two months, she has been slogging away in the field collecting data for her team’s project on human-modified forests. But this year hasn’t gone to plan. This isn’t a case of bad planning though, as with so may projects – 7 of her 20 sites had burned in some of the most widespread fires in recent times. After seeing her tweeting about this, I thought “I need to write something about this. It feels important.” So we fired a few tweets and emails back and forth, with Erika fitting answering my questions between her days in the field. After I had waited impatiently for a couple of days, Erika messaged me:

“Sorry, trying not to work weekends…not going very well though…Today I just learned that 9 of my 20 plots have burned.” 2 more plots. Aside from the wider situation, this was the stuff of researcher’s nightmares.

Erika de Berengeur Cesar in one of her logged plots before it burned (top) and the same plot after recent fires (bottom). Photo courtesy of Erika.

Fires in Brazil reached record levels in 2015, with more than a quarter of a million separate fires recorded. However, these fires are not generally ‘natural’ – “Fires in the region always have a human ignition source.” Erika told me “They are used in slash-and-burn agriculture, to clear pastures of weeds and also to burn downed timber in newly deforested areas.” This year’s strong El Niño has caused drier conditions than normal making it “easier for agricultural fires to escape the targeted area and sweep through the forests.” Indonesia is facing a similar problem, where forests have been burned to clear space for new oil plantations, in what the Guardian’s George Monbiot  has described as the ‘greatest environmental disaster of the 21st century – so far.’

When I queried why it matters that the forest is burning, Erika was clear what the major issue is – the loss of unique biodiversity. “Every year over 100 new species are found in Amazonian forests. To see all this going up in smoke is a crime against humanity. It is a tragedy.”

“How are these fires likely to affect biodiversity?” I asked.

“The Amazon has not co-evolved with periodic fires…This means that Amazonian forests are not used to these events and…do not cope very well with it. In terms of plant communities, there is a sharp increase in the abundance of pioneer species, while high-wood density climax species disappear….Fires negatively affect…rare bird species, and the habitat specialists, such as the ant-following insectivores and the terrestrial gleaners. Overall, burned forests are significantly less diverse than their unburned counterparts.”

Amazonian forests that have burned repeatedly may eventually come to resemble more open savannahs and contain  very different species to relatively undisturbed old-growth forest.

But it’s not just biodiversity that is affected by these fires, but humans as well. In Indonesia there were evacuations of children by the navy, although some of the children, according to reports, still died from breathing difficulties . In Brazil the fires have “affected many of the local people…who reported a number of respiratory problems, such as dry cough, difficulties in breathing, and sore throats,” according to Erika. “People had to spend days building fire breaks to protect their land, instead of directly working on their crops.” People working on these farms already have a tough life as it is, without having to worry if their source of income will go up in smoke.

So what will happen to these forests in the future? Given time and, vitally, protection they can recover but Erika thinks this is unlikely “These burned forests may never recover. After the fire, several large trees die, creating a number of gaps in the forest canopy, through which more light and wind can reach the forest floor, making it drier and, as a consequence, more vulnerable to further fire events.”

The research Erika and her team are carrying out will help to answer the question of how burned forests recover but it is obvious that degraded forests, such as these, need to be seen as a greater conservation priority. More than 50% of the globe’s forests are degraded in one way or another. We cannot afford to only protect primary forests anymore.

Edit: I got an email from Erika a bit ago after I asked her what the best solution would be. I thought I should include it here:

“Funnily enough there are already quite a few good policies in place. The problem is that none is followed. For example, every year there is a ‘burning calendar’ establishing when farmers can use fire to burn their pastures or their croplands. During the peak of the dry season, the use of fire is forbidden. In 2015, given the extreme drought, some states even extended the prohibitive period. So all quite reasonable and good, right? The problem is that no one follows this rules and there is no law enforcement in place. So people carry business as usual and the forests carry on burning. To put in practice the existing laws would be the best solution.”

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If you want to read more about the situation in Brazil take a look at the excellent article Erika has written  for ‘The Conversation.’

There are also a pair of videos that Erika’s team have made documenting the fires that you can see here and here.

Friday linksfest: Messiness and using Google Street View to map species

Anyone that knows me in real life will know that I’m I hate mess. I hate wires that tangle everything up in my flat, creating impenetrable black and off-white thickets. I hated my old office mate leaving his fieldwork kit strewn all across the office, grass festering away on the edges of his quadrat. I hate my sister’s inability to cook without the whole kitchen descending into culinary chaos. In summary, I hate mess.*

So it was with some concern that I saw the recent TED talk by Tim Harford, talking about how messiness can improve problem-solving flash up on my twitter feed. I clicked the link, despairing that I might have to scatter papers all over my desk in an effort to become a better scientist. Thankfully, that wasn’t what happened. Harford actually says that putting constraints or introducing randomness into how you solve problems can mean you produce better solutions. It’s very TED-y, linking disparate ideas from music and science, but I think Harford has something. Give it a look.

 

Something else that grabbed my attention this week was a paper which used Google Street View to build up a picture of where a plant species in Spain can be found. I like papers like this that just blow my mind. I would never have thought differently enough to come up with this idea.

I love using ggplot for analysis in R but up until know I have had to use a different package to do my diagnostic plots for models. Thankfully, now this has been solved with the ggfortify package so you never need to use ugly base R plots again….

And finally, if you are an early career researcher in the UK you might be interested in a joint BES & ZSL meeting aimed at helping you establish a career in conservation. Check it out, it has great speakers and I’m sure it will be as good as all other BES meetings I’ve been to.

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* I do however, love all of the people I mention here. Even if I don’t like the mess you cause.

Beta-diversity – What is it good for?

A while ago I wrote a post asking whether everyone’s favourite measure of biodiversity, species richness, was useful. In it, I concluded that it is probably one of the bluntest, least informative measures of ecological communities we have and that we should try to use alternative metrics when possible. Recently, I started wondering about what other measures of biodiversity might be informative, and what they can be used for. And then a neat review of beta-diversity by James Jacob Socolar ( correction courtesy of James Gilroy on Twitter – thanks James!) and colleagues came out in Trends in Ecology and Evolution, so today I’ll focus on that, borrowing from some of their thoughts and hopefully adding some of my own along the way. In the future, at some point, I’ll write something about temporal changes in ecological communities at individual sites.

So, firstly what do I mean by beta-diversity? Beta-diversity broadly reflects the differences in community composition between sites.  Gamma diversity (regional diversity) is a product of both beta- and alpha-diversity (diversity at a single site). And there are lots* of different ways of measuring beta-diversity. The simplest metric for beta-diversity is termed ‘true beta-diversity’ and was defined by Whittaker in 1960 as:

This metric is perhaps the easiest to interpret, but it also needs a reliable estimate of gamma diversity, so may be difficult to use in practice. Using this method allows the relationship between alpha and gamma diversity to be investigated. Other measures can be based on dissimilarity matrices, identifying pairwise differences between sites. These metrics can then be used to look at drivers of these differences, such as the geographic distance between individual sites and environmental differences. However, dissimilarity matrix methods don’t allow the relationship between alpha and gamma diversity to be investigated. The above explanation probably explains the ubiquity of species richness as a metric in ecology – we can all (more-or-less) agree on what it means.

Changes in beta-diversity when humans alter natural landscapes can be unpredictable. When human disturbances are patchy, such as in the case of selective logging, beta diversity has been shown to be stable or increase due to an influx of generalist species in forest gaps.

Differences changes in tree community dissimilarity with increasing distance between sites in unlogged and logged forest. Note that logged forests show a more rapid rate of change, suggesting that logging results in more variable ecological communities. Figure modified from Berry et al 2008.

In contrast, when human land-use change results in the conversion of natural ecosystems to a relatively homogeneous system in which only a small subset of species can survive, beta-diversity tends to decrease. Examples of such drivers include agricultural conversion and urbanisation. However, even high intensity farming can result in an increase in beta-diversity particularly if species populations decrease leading to greater dissimilarity purely as a result of random processes.  In summary, the response of beta-diversity to anthropogenic change appears to be relatively idiosyncratic.

All of this is well and good, but what use is beta-diversity to practical conservation? At first inspection, this is not clear. The general perception of species richness is that more species = better**. Does higher beta-diversity = better? Well, no, not necessarily. Given that the aims of conservation vary from place to place, it is not surprising that how beta-diversity can be used also varies.

The most obvious use of beta-diversity is in spatial planning of protected areas. In landscapes which show a high spatial turnover of species, managers might favour the use numerous distinct reserves to capture this variation. However, in a landscape in which beta-diversity results from differents in species richness a single protected area might be favoured. Also, if a natural ecosystem is particularly distinct from other candidate sites it may be considered a priority for protection.

High beta-diversity can also result from dispersal limitation in a landscape. For example, secondary forests in fragmented landscapes plants with seeds dispersed by wind may colonise sites more readily than those dispersed by animals that may not cross non-forest areas. So in cases where beta-diversity amongst patches of a similar habitat in a fragmented landscape is high, this may point to the need for restoration to increase connectivity. Successful restoration may result in a decrease of beta-diversity as dispersal between patches increases. For example, Renata Pardini’s work has shown that the small mammal communities of more highly connected fragments of Atlantic forest are more similar to other patches than unconnected fragments. However, as far as I know, there is relatively little evidence empirical that restoration has similar effects.

In the paper I mentioned earlier, Jacob Socolar and colleagues suggest that beta-diversity may also be useful in informing the land-sharing vs land-sparing debate (which i have previously written about here, here an here). They argue that the use of beta-diversity as part of this debate may show that heterogenous landscapes that include agri-environment schemes, management of natural systems and high intensity agriculture are better at maintaining alpha- beta and gamma-diversity. Thus, the incorporation of metrics other than population sizes of species, the classic approach for such comparisons, may produce different conclusions to current studies, which largely suggest land-sparing as a favoured approach. As always with conservation, this depends on what you think we should try to protect. Should we focus on particular species? Or should we look attempt to conserve the processes that maintain coarse-scale diversity?

For me, the key point that the paper makes is that even though two recent high-profie studies recently suggested local-scale alpha-diversity is relatively constant***, global scale gamma-diversity is declining. This suggests that rare species are getting rarer and common species are increasing in abundance. If we can work out how and why beta-diversity responds to land-use changes we can better understand how to conserve gamma-diversity. However, before we do that we need to develop methods to upscale from alpha to gamma diversity and determine how different disturbances alter beta-diversity. Novel approaches offer the potential to solve this problem, but substantial testing is needed to determine how useful they are.

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*Patricia Koleff identified 24 metrics for use with presence-absence data and my  old CEH office mate Louise Barwell tested 29 different beta-diversity metrics that incorporated abundance data. Give both of these papers a read, they’re well worth your time.

**I don’t agree with this perception, I’m just extrapolating based on things I have heard from a few people. Deeply unscientific, I know.

***I saw Andrew Gonzalez present some work on the problems of these two studies at the 2015 British Ecological Society annual meeting and hope to post something when the paper comes out. I can’t say much, but it was fascinating stuff.

 

New paper: Stand dieback and collapse in a temperate forest and its impact on forest structure and biodiversity

We recently published the first paper from my post-doc in Forest Ecology and Management, so I thought I’d share it here. It marks a bit of shift away from the tropical forests I have previously published about (see posts on that here and here), but it allowed me to continue my work on post-disturbance recovery.

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Scientists and policymakers around the world are concerned about the potential effects of forest dieback. Drought and the spread of new pathogens and pests have resulted in increased tree mortality in both the USA and Canada, and these threats are likely to increase in Europe as well. The IPCC recently highlighted forest dieback as a potential major threat, but one about which we know relatively little.  Changes in forest biodiversity and ecosystem services are likely to be particularly severe in ecosystems that show poor resilience. Failure to withstand or recover from drought or pest attack may lead to ‘regime shifts’ resulting in a very different type of system, with many fewer trees.

Die back in the New Forest has caused some areas to go from this…

…to this

Luckily for our group my boss, Adrian Newton, found out about a permanent transect that had been set up in the 1950s in a woodland in the New Forest that now appears to be suffering from dieback. The site had been surveyed 4 times between 1964 and 1999, and our team collected more data from the site in 2014. In our recent paper, published in Forest Ecology and Management we used this data to investigate dynamics of the woodland. In particular, we addressed the potential impacts of dieback on forest structure, the causes of these changes and their impact on biodiversity.

Basal area loss in Denny wood from 1964-2014

To cut a long story short, the forest lost about a third of its basal area (as you can see above) and over two-thirds of its juvenile trees over 50 years. over 90% of the loss of basal area was due to the death of large beech (Fagus sylvatica) and oak (Quercus rubor) trees.

Climate records from 1964-2014 showed that (a) mean temperature during April-September increased from 1960s to present day; and (b) there were numerous drought years post-1976 a year which was previously identified as a cause of current mortality.

The external factors causing these changes are not entirely clear, but there have been a number of significant droughts between 1964-2014 as well as increased temperatures (see figures above). In addition, the presence of a number of novel fungal pathogens has been noted in the forest, which may have interacted with drought to further weaken large trees. Recovery in the forest has been very limited, with almost no recruitment of saplings of the canopy dominants (beech & oak) in 50 years. This low recruitment is probably a result of the high density of ponies and deer in the woodland.

Relationship between percentage loss in subplot basal area and (a) percentage grass cover and (b) ground flora species richness.

The result of the changes in forest structure is that areas with little tree cover have seen large increases in grass cover and increased ground flora species richness (see figure above). Both of these results indicate that there may be a tipping point at which changes in structure result in rapid increases in grass cover and species richness of ground flora.

Many of the papers on resilience talk about alternative stable states, in which transitions from one type of system to another are difficult to reverse. Though, from the outside, it may appear that our field site shows evidence of a shift to a relatively treeless stable state, we think that this is incorrect. The theory underlying multiple stable states suggests that disturbances causing the regime shift should be a ‘pulse’, when disturbance occurs over a relatively short period and then does not occur again, rather than a ‘press’ disturbance, where the disturbance is present over long periods of time.  However, these conditions are not met by our site where both pulse (i.e. drought) and ongoing press (i.e. overgrazing)  disturbances are present. We think that both of these processes are needed to cause the forest to lose tree cover.

Even if the transition we  have observed is not strictly a ‘regime shift’ it’s still important. Dieback is apparently widespread in the New Forest and is on-going, so the potential impacts could be very significant. As with other cases of dieback it’s difficult to identify appropriate management responses. However, in the case of the New Forest the easiest way to restore resilience would be to protect tree regeneration from the high herbivore pressure in the area.

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If you want to read more about our study you can find the paper here and details of our project on forest resilience can be found here. Oh, and here’s a post I wrote about my project a while back. Also, feel free to comment below!